Biodegradation of fluorinated compounds widely used in agro-industrial applications

Diogo Alves da Mota Alexandrino

M 2016

DISSERTAÇÃO DE MESTRADO

TOXICOLOGIA E CONTAMINAÇÃO AMBIENTAIS

Diogo Alves da Mota Alexandrino

BIODEGRADATION OF FLUORINATED COMPOUNDS WIDELY

USED IN AGRO-INDUSTRIAL CONTEXTS

Dissertação de Candidatura ao grau de

Mestre em Toxicologia e Contaminação

Ambientais submetida ao Instituto de

Ciências Biomédicas de Abel Salazar da

Universidade do Porto.

Orientadora – Doutora Maria de Fátima

Carvalho

Categoria – Investigadora Auxiliar

Afiliação – Centro Interdisciplinar de

Investigação Marinha e Ambiental da

Universidade do Porto

Co-orientadora – Doutora Ana Paula Mucha

Categoria – Investigadora Auxiliar

Afiliação – Centro Interdisciplinar de

Investigação Marinha e Ambiental da

Universidade do Porto

ACKNOWLEDGEMENTS

Firstly, I would like to thank my supervisor, Dr. Maria F. Carvalho, without whom

the work integrated in this thesis would have not been possible. I genuinely

thank her incredible dedication and trust, certain that part of my future goals

have been established as a result of her mentoring, which became both an

incredible honour and a fundamental phase in my personal and professional

development.

I would also like to thank my co-supervisor, Dr. Ana Paula Mucha, to whom I

thank for the opportunity of integrating her laboratory, where I was always given

all the conditions to develop my work to the fullest of its potential.

Secondly, I would like to acknowledge CIIMAR - Interdisciplinary Centre of Marine

and Environmental Research and Departamento de Química e Bioquímica of

Faculty of Sciences of University of Porto, for the use of all the equipment,

installations and facilities.

To my lab mates at Ecobiotec (CIIMAR-UP), I recognise their friendship, as well

as all the input in my work and precious help and support, with a special

emphasis to Patricia Duarte, Filipa Santos, Joana Fernandes and Inês Ribeiro.

I am also grateful for the involvement of Dr. Marisa Almeida, Dr. Filipe Pereira

and Prof. Rui Oliveira in my work: to Dr. Marisa Almeida, I thank her support in

all HPLC analysis, as well as the input in the scientific revision in one of the

submitted scientific manuscripts that are part of this thesis; to Dr. Filipe Pereira

for his incredible support and input regarding the molecular biology tools used

in this work; to Prof. Rui Oliveira for his mentorship since my Bachelor Degree

and for aiding on the revision of part of this thesis.

Finally, I would like to give my deepest regards to my family and close friends,

acknowledging their friendship, support and company during this stage of my

life. Especially to my parents, I thank them for always enabling and supporting

me, shaping the person I am today.

The research contemplated in this thesis was supported by the Strategic Funding

UID/Multi/04423/2013 through national funds provided by FCT – Foundation

for Science and Technology and European Regional Development Fund (ERDF),

in the framework of the programme PT2020, by the structured Program of R&D&I

INNOVMAR - Innovation and Sustainability in the Management and Exploitation

of Marine Resources, reference NORTE-01-0145-FEDER-000035, namely within

the Research Line ECOSERVICES (Assessing the environmental quality,

vulnerability and risks for the sustainable management of the NW coast natural

resources and ecosystem services in a changing world) within the R&D Institution

CIIMAR (Interdisciplinary Centre of Marine and Environmental Research),

supported by the Northern Regional Operational Programme (NORTE2020),

through the European Regional Development Fund (ERDF), and by Investigador

FCT program supported by FCT, FSE and Programa Operacional Potencial

Humano, within the scope of the project IF/00791/2013/CP1197/CT0002.

ABSTRACT

Fluoroorganic compounds are a class of chemicals that are thriving in virtually

all economic sectors, essentially due to the unique properties of the fluorine

atom. The pharmaceutical and agrochemical industries are two important

sectors where these compounds are used, with a wide range of commercial

drugs and pesticides belonging to this class of compounds. The aim of this

thesis was to investigate the biodegradation of fluoroorganics with distinct

chemical structures (aliphatic and aromatic) and applications.

In the first experimental work, the biodegradation of a group of structurally

related aliphatic carboxylic fluoroorganics – mono- (MFA), di- (DFA) and

trifluoroacetate (TFA) - was investigated, using a variety of environmental

samples as a microbial source. Biodegradation experiments were carried out

under different modes of substrate supplementation, which included (i)

fluoroacetates (FAs) fed as sole carbon source; (ii) FAs (only for DFA and TFA)

fed in co-metabolism with sodium acetate and (iii) mixtures of MFA with DFA or

TFA. Biodegradation of the target compounds was assessed through fluoride ion

release. The results obtained revealed that from the three FAs fed, only MFA was

completely defluorinated, while DFA and TFA were recalcitrant in all tested

conditions. When present in mixture, DFA was shown to inhibit biodegradation

of MFA, whereas TFA had no effect. A total of 15 bacterial isolates were found

to degrade as single strains 20 mg L-1

of MFA as sole carbon source. 16S rRNA

gene sequencing analysis indicated that from these degrading bacteria, only

Delftia acidovorans had been previously reported to degrade MFA. This work

shows that biodegradation of the three tested FAs is very distinct, despite these

compounds being structurally related, and draws the attention to the unknown

impacts that the accumulation of DFA and TFA may have in the environment as

a result of their high recalcitrance.

In the second experiment, biodegradation of a veterinary antibiotic, enrofloxacin

(ENR), was investigated both individually and in mixture with a non-fluorinated

antibiotic, ceftiofur (CEF). Biodegradation was investigated for a concentration

range between 1-3 mg L-1

and using acetate as a co-substrate. Microbial inocula

were obtained from rhizosphere sediments of plants derived from experimental

constructed wetlands designed for the treatment of livestock wastewaters

contaminated with trace amounts of these antibiotics. Complete removal of CEF

from the inoculated culture medium was always observed, independently of its

concentration or the concomitant presence of ENR. Biodegradation of ENR

decreased with the increase in its concentration in the culture medium, with

defluorination percentages decreasing from ca. 80 to 4 % in the cultures fed with

1 and 3mg L-1

, respectively. Ciprofloxacin and norfloxacin were detected as

biodegradation intermediates of ENR degradation in the inoculated culture

medium supplemented with this antibiotic, indicating that defluorination of at

least part of ENR in these cultures is not an immediate catabolic step. Abiotic

mechanisms showed to have a high influence in the removal of CEF, affecting

less ENR degradation. The enrichment process with the target antibiotics led to

significant shifts in the structure and diversity of the microbial communities,

predominantly selecting microorganisms belonging to the phyla Proteobacteria

(e.g. genera Achromobacter, Variovorax and Stenotrophomonas) and

Bacteroidetes (e.g. genera Dysgonomonas, Flavobacterium and

Chryseobacterium). The results presented in this study indicate that

biodegradation can be an important mechanism for the environmental removal

of the tested compounds. In overall, the two developed works indicate that

fluorinated compounds are a challenge for microbial degradation yet, due to the

high metabolic versatility of microorganisms, biodegradation is still a possible

mechanism for their environmental remediation. The results obtained in the

present thesis also indicate that the degree of fluorination and compound

concentration have a crucial role in the recalcitrance of fluorinated compounds.

RESUMO

Os compostos organofluorados constituem uma classe de compostos químicos cuja

utilização se encontra em expansão em praticamente todos os setores económicos,

essencialmente devido às propriedades únicas do átomo de flúor. Os setores

agroquímico e farmacêutico constituem dois segmentos industriais onde esta classe

de compostos tem especial relevância, dado o elevado número de produtos

farmacêuticos e pesticidas fluorados atualmente comercializados. O objetivo desta

dissertação foi investigar a biodegradação de compostos orgânicos fluorados com

distintas estruturas químicas (alifáticos e aromáticos) e aplicações práticas.

No primeiro trabalho experimental investigou-se a biodegradação de três

compostos fluorados alifáticos estruturalmente semelhantes – mono- (MFA), di-

(DFA) e trifluoroacetato (TFA) – utilizando como inóculos, microrganismos

provenientes de diferentes amostras ambientais. Nas experiências de

biodegradação realizadas, os fluoroacetatos (FAs) foram suplementados de

diferentes modos: (i) FAs como fonte única de carbono; (ii) DFA ou TFA em

cometabolismo com acetato e (iii) misturas de MFA com DFA ou TFA. A libertação

do ião fluoreto foi utilizada como indicador da biodegradação dos FAs. Os

resultados obtidos revelaram que dos três FAs alimentados apenas o MFA foi

completamente defluorinado, enquanto o DFA e TFA foram recalcitrantes em todas

as condições testadas. Quando em mistura, a presença de DFA inibiu a

biodegradação de MFA, enquanto o TFA não teve qualquer efeito inibitório. Um total

de 15 isolados bacterianos mostraram ser capazes de degradar individualmente 20

mg L-1

de MFA como fonte única de carbono. A sequenciação do gene 16S rRNA

desses isolados revelou que apenas a espécie Delftia acidovorans foi anteriormente

reportada como degradadora de MFA. Estes resultados mostram que a

biodegradação destes três FAs é bastante distinta, apesar das suas similaridades

estruturais, e chamam a atenção para a importância de conhecer os impactos

decorrentes da persistência e acumulação de DFA e TFA no ambiente, como

resultado da elevada recalcitrância destes compostos.

Na segunda experiência, estudou-se a biodegradação de um antibiótico veterinário,

enrofloxacina (ENR), suplementado individualmente e em mistura com um

antibiótico não fluorado, ceftiofur (CEF). A biodegradação foi investigada para uma

gama de concentrações entre 1-3 mg L-1

, utilizando acetato como co-substrato.

Utilizou-se como inóculo, rizosedimento de plantas provenientes de uma fito-etar

experimental desenhada para o tratamento de efluentes de pecuária contaminados

com concentrações vestigiais dos antibióticos estudados. A completa remoção de

CEF foi sempre observada, independentemente da sua concentração nas culturas

microbianas ou da concomitante presença de ENR. A biodegradação de ENR

diminuiu com o aumento da sua concentração no meio de cultura, com percentagens

de defluorinação oscilando entre os 80 e os 4 % nas culturas suplementadas com 1

e 3 mg L-1

, respetivamente. Os intermediários metabólicos ciprofloxacina e

norfloxacina foram detetados nas culturas suplementadas com ENR, indicando que

pelos menos parte da molécula de ENR não é imediatamente sujeita a uma reação

de defluorinação. Os mecanismos abióticos mostraram ter uma grande influência na

remoção de CEF, não afetando de forma tão acentuada a degradação de ENR. O

processo de enriquecimento com os antibióticos estudados levou a alterações

significativas ao nível da estrutura e diversidade das comunidades microbianas,

selecionando predominantemente microrganismos pertencentes aos filos

Proteobacteria (p. ex. géneros Achromobacter, Variovorax e Stenotrophomonas) e

Bacteroidetes (p. ex. géneros Dysgonomonas, Flavobacterium e Chryseobacterium).

Os resultados deste estudo mostraram que a biodegradação pode ser um

importante mecanismo na remoção destes antibióticos do ambiente.

De uma forma geral, ambos os trabalhos realizados mostram que os compostos

orgânicos fluorados constituem um desafio para a degradação microbiana, no

entanto, atendendo à elevada versatilidade metabólica dos microrganismos, a

biodegradação destes compostos revela-se um mecanismo viável para a sua

remediação ambiental. Os resultados obtidos indicam também que tanto o grau de

fluorinação como a concentração do composto têm um papel fundamental na

recalcitrância dos compostos fluorados.

TABLE OF CONTENTS

Abbreviations and Syncronims I

List of Figures II

List of Tables III

CHAPTER 1- Introduction

1. Xenobiotics in the environment 1

2. Fluoroorganic compounds

2.1. Elemental fluorine 3

2.2. The C-F bond 3

2.3. Biological significance of fluoroorganics 5

2.4. Industrial significance of fluoroorganics 6

3. Biodegradation and Bioremediation

3.1. Factors influencing biodegradation and

bioremediation 8

4. Microbial transformation and degradation of fluoroorganic

compounds 11

5. Genomic and metagenomic approaches in biodegradation

studies 13

6. Aim and outline of this thesis 14

CHAPTER 2 – Biodegradation of Mono-, Di- and Trifluoroacetate by Microbial

Inocula with Different Origins

1. Introduction 17

2. Materials and Methods

2.1. Microbial inocula 19

2.2. Biodegradation experiments 19

2.3. Bacterial characterization of MFA-degrading

cultures 20

2.4. Biodegradation capacity of bacterial isolates

obtained from MFA-degrading cultures 20

2.5. Identification of MFA-degrading isolates 21

2.6. Analytical methods 22

3. Results

3.1. Biodegradation of FAs by the different microbial

inocula 22

3.2. Characterization of MFA-degrading bacterial

consortia and biodegradation capacity of the isolated

strains 24

3.3. Identification of the MFA-degrading bacterial

isolates 26

4. Discussion 26

5. Conclusion 32

CHAPTER 3 – Biodegradation of the Veterinary Antibiotics Enrofloxacin and

Ceftiofur and Associated Microbial Community Dynamics

1. Introduction 34

2. Materials and Methods

2.1. Enrichment of microbial degrading cultures 35

2.2. Biodegradation of different concentrations of ENR

and CEF 36

2.3. Analytical methods 37

2.4. Analysis of the structure of the microbial

communities 38

2.5. Statistical analysis 39

3. Results

3.1. Biodegradation of ENR and CEF 40

3.2. Analysis of microbial communities’ dynamics 43

4. Discussion 44

5. Conclusion 52

CHAPTER 4 – General Discussion and Conclusions

1. General discussion 55

2. Conclusion 57

CHAPTER 5 – References 60

I

ABBREVIATIONS AND SYNCRONIMS

Å - Ångstrom

Bp – Base pair

CEF – Ceftiofur

C-F – Carbon-fluorine

CIP – Ciprofloxacin

CP – Cephalosporins

DFA - Difluoroacetate

ENR – Enrofloxacin

FAdH – Fluoroacetate dehalogenase

FAs - Fluoroacetates

FQ – Fluoroquinolones

HCFC - Hydrochlorofluorocarbon

HFC – Hydrofluorocarbon

Kj – Kilojoule

LOD – Analytical Limit of Detection

LOQ – Analytical Limit of Quantification

MFA - Monofluoroacetate

MM – Minimal medium

NOR – Norfloxacin

OD – Optical Density

OTU – Operational Taxonomic Unit

PCA – Plate-Count Agar

PCR – Polymerase Chain Reaction

Pm - Picometer

QIIME – Quantitative Insights into Microbial Ecology

rRNA – Ribosomal RNA

STE – Sodium Chloride-Tris-EDTA

TFA - Trifluoroacetate

TISAB III – Total Ionic Strength Adjusting Buffer

WWTP – Wastewater Treatment Plant

II

LIST OF FIGURES

Figure 1 Main anthropogenic sources of organofluorine contaminants

and their corresponding environmental dynamics. 2

Figure 2 Chemical structure of the four top-selling fluorinated

pharmaceuticals in 2015. 5

Figure 3 Biodegradation performances, based on fluoride release, of

MFA supplemented as a sole carbon source during a two

months period.

23

Figure 4 Defluorination performance of the tested microbial consortia

when supplemented with mixtures of FAs after two feeding

periods.

25

Figure 5

Biodegradation based on fluoride release of FAs supplemented

as sole carbon sources and in co-metabolism with MFA, by a

mix of the 13 MFA-degrading microbial isolates.

28

Figure 6 Biodegradation of ENR, supplied individually and in a mixture

with CEF for the concentrations of 3 and 2 mg L-1

. 42

Figure 7 Removals of ENR and CEF obtained in different experimental

conditions, for the concentration of 2 mg L-1

. 43

Figure 8 Cluster analysis based on Bray-Curtis similarity of

metagenomics profiles of microbial communities and relative

abundance of the different bacterial phyla at the beginning and

at the end of the biodegradation experiments.

46

III

LIST OF TABLES

Table 1 Physicochemical properties of the different halogens 4

Table 2 Microbial strains isolated from the different MFA-degrading

consortia and taxonomic identification of the microbial

isolates capable of degrading MFA as a sole carbon source

27

Table 3 Defluorination performance along a feeding period of 21 days,

obtained nine weeks after the beginning of the enrichment

phase, for ENR supplied individually and in mixture with CEF,

at the concentration of 1 mg L-1

41

Table 4 Diversity and abundance indexes of the initial inocula and

microbial communities enriched with the target antibiotics

45

Table 5

Metagenomics profiles of the initial inocula and enriched

consortia, showing the relative abundance of each taxonomic

group in the communities

47

1 CHAPTER

INTRODUCTION

1

1. Xenobiotics in the environment

The advances in chemical synthesis have led to the introduction of countless new

compounds in all segments of contemporary societies and to the generation of

novel products and materials. Today, more than ever, the creation, production,

marketing and overall use of novel synthetic and semi-synthetic products happens

in an unequalled rate.

While industrialization has been the main drive towards the high standards of living

that exist nowadays, this phenomenon is also the main responsible for the

environmentally-threatened Earth that we live today. As a result of the rapid

industrial development that occurred after the first Industrial Revolution, which was

accompanied by a hasty urbanization and an increase on world population, a

significant anthropogenic pressure in every component of the environment has

been occurring.

The production, introduction and spreading of xenobiotics in the environment is a

direct cause of the increased anthropogenic footprint in the environment. Being

compounds foreign to nature, xenobiotics have an increased potential for

ecosystem damage, attending to their capability of disrupting the dynamics of

nature. Moreover, the recycling and natural removal of such products from

environmental matrices is not always possible due to their foreign nature and

constant environmental input.

2. Fluoroorganic compounds

The first reported synthesis of a fluoroorganic compound dates back to mid-18th

century, but it was only in the 1930s that these products gained an industrial

dimension with the production of chlorofluorocarbons and other industrially

relevant fluorinated products (Okazoe, 2009; Kirsch, 2013). With the development

of new methodologies enabling a more efficient synthesis of the carbon-fluorine

(C-F) bond, the overall manufacturing of synthetic organofluorines skyrocketed.

In addition to industrial applications, the unique properties of fluoroorganics made

these compounds also attractive for other types of applications. For example, the

discovery of the first fluorinated pharmaceutical by Heidelberger et al. (1957) (5-

fluorouracil, an anticancer drug) drew attention to the role that fluoroorganic

compounds could have on the design of pharmaceuticals and agrochemicals.

2

Nowadays, this class of compounds has a widespread use in various applications,

ranging from pharmaceuticals, agrochemicals and biocides, industrial reagents,

solvents, anti-adherents, plastics, fire retardants, refrigerants, anaesthetics, among

others (Key et al., 1997; Kiel and Engesser, 2015). As a result of this growing use

in most economical sectors, the environmental presence of organofluorine

compounds has witnessed a proportional increase (Fig. 1) (Key et al., 1997; Kiel

and Engesser, 2015).

Biological production of fluorinated molecules is very rare in nature. Biogenic

halogenation is verified in ca. 3700 organic molecules, but only about 20 of these

correspond to fluorinated structures (Gribble, 2003; Kiel and Engesser, 2015). All

the known naturally-produced organofluorines are monofluorinated, which

contrasts with synthetic fluoroorganic compounds which usually have more than

one fluorine atom in their molecules (Key et al., 1997). Thus, the overwhelming

majority of organofluorinated compounds are foreign to nature, being xenobiotics,

by definition. In addition, the scarce occurrence of natural fluorinated structures

indicates that fluorinated compounds do not have a central role in biological

processes (Kirsch, 2013).

Figure 1. Main anthropogenic sources of organofluorine contaminants and their corresponding

environmental dynamics.

3

2.1. Elemental fluorine

Fluorine is a chemical element belonging to the halogens group. It was first

discovered in 1810 by André-Marie Ampére, but the isolation of this element was

only achieved in 1886 by Henry Moissan (Chang, 2010). Despite its early discovery,

fluorine only became relevant almost a century later (Okazoe, 2009).

Although all halogens are highly reactive, fluorine exhibits an unprecedented

reactivity, as it is highly oxidizing and prone to radical formation (Jaccaud et al.,

2000; Kirsch, 2013). Also, this element has an extreme electronegativity (and,

consequently, a very high ionization energy), which further contributes to its high

reactivity (Chang, 2010; Kirsch, 2013).

Due to its peculiar properties and widespread uses, fluorine has been

acknowledged has the “small atom with a big ego” (Uneyama, 2007). In fact, this

element has unique physicochemical properties (Table 1), which justify its current

diversified applications. However, these properties also render all fluorinated

molecules – either organic or inorganic – a certain outlandishness in terms of

structure, reactivity and overall biotic and abiotic behaviour.

The ionic form of fluorine, fluoride, has a very small ionic radius (Table 1), similar

to a hydroxyl anion or a hydrogen, meaning that the replacement of a hydrogen

atom or a hydroxyl group by fluorine occurs with minimal steric interferences

(Jaccaud et al., 2000). Also, fluorine is capable of establishing with carbon one of

the strongest chemical bonds known in organic chemistry.

When compared to the other halogens, fluorine is the most abundant, being also

one of the most common elements in the planet (Jaccaud et al., 2000). However,

this element occurs mainly in inorganic forms, integrating various minerals

(fluorspar, fluorite, fluorapatite, cryolith and topaz) (Harnisch and Eisenhauer,

1998). In fact, the natural occurrence of fluorine embedded in organic molecules

is a very rare phenomenon.

2.2. The C-F bond

The peculiar properties of organofluorinated compounds can be partially attributed

to the special nature of the chemical bond that fluorine establishes with carbon in

organic molecules.

4

This bond is thought to be one of the strongest in organic chemistry, partially due

to the high electrostatic attraction between fluorine and carbon and to the excellent

orbital compatibility between these two elements (Banks et al., 1994; O'Hagan,

2008; Kirsch, 2013). In addition, due to the extreme electronegativity of the

fluorine atom, when fluorine is bonded to carbon it always attracts more strongly

the shared electrons, creating a highly polarized chemical bond.

Table 1. Physicochemical properties of the different halogens

Property Fluorine Chlorine Bromine Iodine Ref.

Melting point (ºC) -223 -102 -7 114 Chang (2010)

Boiling point (ºC) -187 -35 59 184 Chang (2010)

Atomic radius (pm) 72 99 114 140 Chang (2010)

Ionic radius (pm) 133 181 196 216 Haynes (2014)

Ionization energy (kJ mol-1

) 1680 1251 1139 1009 Chang (2010)

Electronegativity (Pauling Scale) 4.0 3.0 2.8 2.7 Chang (2010)

Bond strength when bounded to

carbon (kJ mol-1

)

485 339 285 213 Banks et al. (1994)

Besides, the C-F bond also has a small length (1.35 Å), being only compared to

carbon-hydrogen (1.09 Å) and carbon-oxygen bonds (1.43 Å). In fact, the C-F bond

possibly represents the smallest chemical bond between carbon and a heteroatom

in organic molecules.

As a result of the properties of the C-F bond, fluorinated molecules are less likely

to interact with neighbouring molecules (Murray-Rust et al., 1983), with this

property closely influencing pharmacokinetics and environmental dynamics of

these compounds.

5

2.3. Biological significance of fluoroorganics

Fluorine-substituted molecules have a high potential as biologically active

compounds in areas ranging from medicinal chemistry to agriculture (Ojima, 2013).

The applications of organofluorine compounds as therapeutics, diagnostic agents,

pesticides, among others, are rapidly expanding, much due to the special

properties of these compounds. One clear example of this is demonstrated by the

large number of fluorinated pharmaceuticals currently approved for human and

veterinary use (Bégué and Bonnet-Delpon, 2006; Isanbor and O’Hagan, 2006;

Yamazaki et al., 2009). In fact, ca. 25% of pharmaceuticals currently

commercialised correspond to fluorinated compounds and from the ten most sold

human pharmaceuticals in the year of 2015, four of them are fluorinated – Crestor®

(rosuvastatin), Sovaldi® (sofosbuvir), Advair Diskus® (fluticasone propionate) and

Januvia® (sitagliptin), respectively (Fig. 2) (Gilchrist, 2015; Murphy, 2016).

Organofluorine pesticides such as fipronil, epoxiconazole and trifluralin, can also

be found amongst the top best-selling agrochemicals in Europe and in the United

States of America (Loi et al., 2011).

Figure 2. Chemical structure of the four top-selling fluorinated pharmaceuticals in 2015.

6

The reason why fluorine is becoming one of the most attractive heteroatoms in

molecular design, is fully attributed to the atomic properties of this element which

are transferred to the molecules it incorporates. In other words, its peculiar

characteristics are mirrored in the compounds it incorporates, leading to the

emergence of favourable properties.

One of the most significant attributes associated with the molecular incorporation

of fluorine, is the increase in the metabolic stability of fluorinated structures (Zhang

et al., 2012; Ojima, 2013). Due to their molecular strength and low reactivity,

fluoroorganic compounds are likely to remain stable in blood circulation, reducing

their susceptibility to detoxification mechanisms and also their potential for

systemic toxicity (Ojima, 2013). On the other hand, the low likelihood of

intermolecular interactions associated with their increased metabolic stability,

promotes the selectivity of organofluorine compounds. By reducing their

interaction with secondary targets, fluorinated molecules will exert their bioactivity

more directly and efficiently.

Lipophilicity and membrane permeability is significantly promoted with the

incorporation of fluorine in aromatic molecules, and thus fluorinated compounds

have enhanced pharmacokinetics and pharmacodynamics properties (Zhang et al.,

2012; Ojima, 2013). This is a favourable property for both pharmaceuticals and

biocides, as it promotes their biological activity in biochemical and physiological

targets.

As already referred, fluorine shares a similarity in steric size with hydrogen or a

hydroxyl group (Jaccaud et al., 2000). This means that fluorine-substitution can

generate congeners of desirable chemical structures with enhanced characteristics,

without compromising its intended biological effect. The production of synthetic

or semi-synthetic fluorinated compounds is becoming a common trend in medicinal

chemistry, with a special focus on the biosynthesis of fluorinated analogues of

natural products (Zhang et al., 2012).

2.4. Industrial significance of fluoroorganics

Fluorinated compounds hold several properties that are highly attractive for

industrial purposes, where they may act as reagents, solvents, building blocks,

polymers, among others. The physical, chemical and thermal stability of the C-F

bond is the main reason why fluoroorganics are highly used in industry. The

7

distinctive solubility properties induced by molecular fluorination, make some

organofluorine compounds to act as optimal industrial solvents, being compatible

with most lipophilic substances or with other fluorinated compounds. These

solvents are highly used in purification processes in fine-chemistry industries and

in the synthesis of other fluoroorganic products (Spargo, 2005).

Fluorinated compounds may also serve as important building blocks, being

frequently used as fluorine donors or precursors in the preparation of more

complex organofluorines (Siegemund et al., 2000).

Fluorinated polymers have increased advantages, essentially associated with their

high resistance, great isolating properties and non-permeability (Siegemund et al.,

2000). As a result of their versatility, various fluoropolymers are currently being

used in many applications, ranging from domestic appliances, cookware, textiles,

clothing, medical equipment, or even in the formulation of firefighting foams

(Siegemund et al., 2000). One good example of a widely used fluoropolymer is

polytetrafluoroethylene, a waterproof and light polymer that is part of the famous

materials Teflon® and GoreTex®.

3. Biodegradation and Bioremediation

Biodegradation is a biological process carried out by microorganisms that leads to

the simplification of the molecular structure of a compound, as a result of the

catabolic activity of microbial enzymes. Bioremediation, refers to the strategic

employment of microorganisms with the capacity to attenuate a contamination

scenario, either by removing or neutralizing the target contaminants (Crawford,

1998). Both these concepts are closely related, as bioremediation strategies rely on

the biodegradation potential of microorganisms and make use of their capacities

to ensure environmental restoration. During bioremediation, microbial removal or

transformation of xenobiotics into less, or even non-hazardous products occurs

(Karigar and Rao, 2011).

Since the 1980s, bioremediation processes have been employed for the

remediation of oil spills and other hazardous products (Shannon and Unterman,

1993) and, more recently, they have also been applied in different contexts of

contamination, mostly targeting micropollutants and emerging pollutants (Das and

Dash, 2014).

8

Contaminants may also be removed from the environment through

physicochemical processes, such as precipitation, coagulation, adsorption,

biosorption or reverse osmosis (US-EPA, 2007; Wang and Chen, 2009; Das and

Dash, 2014). However, when compared to these processes, bioremediation

presents several advantages: (i) biological remediation of contaminated sites tends

to be cheaper than physicochemical remediation techniques (Kumar et al., 2011);

(ii) some physicochemical processes of remediation are highly invasive and, as a

result, may yield secondary effects in the environment, (iii) bioremediation has the

potential to mineralize the contaminants, i.e., to convert the contaminants into

their constituent elements, as it is based on natural and recycling processes; (iv)

bioremediation technologies are capable of removing environmental contaminants

with minimal environmental impacts and, most of the times, without involving the

transfer of contaminated waste or soil for ex situ treatment, as in many

physicochemical processes (Kumar et al., 2011).

In spite of its several advantages, bioremediation also presents some limitations.

Biological remediation strategies rely on the efficiency of metabolically-competent

microorganisms, however these may not always be present or active at the

bioremediation site. Also, metabolic reactions are always dependent on microbial

viability, which in turn is highly influenced by variables that are hard to control in

real-life scenarios, such as suitable environmental conditions or appropriate levels

of nutrients (Kumar et al., 2011). Moreover, in order to ensure the expression of

the key enzymes involved in the bioremediation processes, several conditions

should be met, including adequate concentration of the target contaminant in the

environment, suitable temperature, pH and redox conditions and suitable

bioavailability of the contaminant. In addition, certain bioremediation strategies

may affect the normal dynamics of ecosystems. For example, the use of non-

autochthonous microbial species/microbial communities for bioremediation

purposes may cause disturbances on the ecology and microbial dynamics of the

indigenous microbiota of the site, impairing the natural functioning of that

ecosystem (Thompson et al., 2005; Kumar et al., 2011).

3.1. Factors influencing biodegradation and bioremediation

Microbial metabolism is a central aspect of biodegradation and bioremediation, as

it is determinant for the transformation and environmental removal of

contaminants. Microorganisms are able to convert or even mineralize xenobiotics

9

through catabolic reactions, usually associated with energy consumption (Adams

et al., 2015). These metabolic processes normally involve redox reactions and may

be associated with respiration or other biological functions that are indispensable

for cell viability and reproduction (Adams et al., 2015). Such reactions are highly

influenced by various factors, either intrinsic to the microorganisms or associated

with the environment where they are integrated, and will directly influence the

overall effectiveness of biodegradation and, consequently, of the bioremediation

strategy.

The capacity of microorganisms to transform, accumulate or mineralize

contaminants is a fundamental part of bioremediation (Karigar and Rao, 2011). In

a contamination scenario, prior knowledge on this aspect is needed for the

outlining of an efficient bioremediation strategy. The capacity of microbial cells to

utilize xenobiotic structures as sources of energy is not a common phenotype,

since their catabolic enzymes did not have a natural evolution process with these

compounds. While microbial metabolism can be extremely versatile, some

contaminants remain recalcitrant to biological degradation, especially when having

high molecular weights and bearing complex ring structures and halogen

substituents (Das and Dash, 2014). Many xenobiotic compounds are biodegraded

only through cometabolic processes. Cometabolism can be defined as a metabolic

interactive effect between two substrates, where usually one is actively

metabolised, being used as a source of carbon and/or energy, and the other one is

unable to support microbial growth (Criddle, 1993). Different variations of

cometabolic reactions might occur in the environment, contributing to the

conversion of various chemical compounds, either by supporting an increase in

microbial density or by improving metabolic performances (Dean-Ross et al., 2002).

The capacity to degrade a certain contaminant or group of contaminants may be

intrinsic to autochthonous microbial species/microbial communities or not. If the

native microbiota of a contaminated site includes microorganisms capable of

metabolizing a contaminant, then the bioremediation strategy can make use of

these microorganisms to remove the pollutants (Das and Dash, 2014). There are

cases though, where the introduction of exogenous, non-native, microbial species

is needed in order to remove specific contaminants (bioaugmentation). In any of

these situations, the microorganisms responsible for the biodegradation processes

must be able to reach their optimal activity and metabolic peak in contaminated

sites, so that they are able to remove or neutralize the xenobiotics. In order to

10

achieve this, in some cases it is necessary to add nutrients, essentially nitrogen and

phosphorous, to the contaminated site (biostimulation).

While knowledge on microbial metabolic potential is a key factor in bioremediation

processes, information on other variables, such as environmental factors or

contamination dynamics is also very important. Environmental factors include a

wide array of physical, chemical and biological conditions that confer additional

complexity to the whole bioremediation process, as they influence both microbial

activity and the environmental dynamics of the contaminants. Among the broad

spectrum of environmental variables the most relevant ones are: geophysical

characteristics of the affected site, nutrient availability, presence of oxygen (or

other electron acceptors), temperature and pH (Das and Dash, 2014; Adams et al.,

2015). Site characteristics should be properly explored prior to the implementation

of a bioremediation strategy. Besides influencing the distribution and

bioavailability of the contaminants, it will also determine microbial survival rate by

modulating oxygen content, nutrient availability, water content, among other

factors (Adams et al., 2015). Nutrients are essential elements for the survival,

viability and multiplication of microbial cells, with carbon, hydrogen and nitrogen

being needed in greater quantities over other elements (Das and Dash, 2014).

Temperature, pH or oxygen content, are vital factors for microbial survival and their

optimal levels will depend significantly on the type of microorganisms involved in

the biodegradation mechanisms.

Regarding the dynamics of contamination, the magnitude, extent, mobility and

toxic potential of the involved contaminants are essential aspects (Das and Dash,

2014). Knowledge on this will allow to clarify the hazardous nature of the

contamination and to more properly define the bioremediation strategy. The

characteristics of the contamination may pose as a limiting factor in bioremediation

– the type of contaminants, as well as their formulation, concentration and

bioavailability will always determine the likelihood of microbial degradation and,

thus, the efficacy of the bioremediation process (Adams et al., 2015). Geological

and soil characteristics of the site are also important in the environmental dynamics

of contaminants, as they influence their mobility, distribution and bioavailability.

11

4. Microbial transformation and degradation of fluoroorganic compounds

The recalcitrant nature of fluoroorganics is widely acknowledged and has been

verified for various fluorinated compounds (Key et al., 1997; Neilson and Allard,

2002). Yet, when concerning the biodegradation of these compounds, scientific

research has focused more on fluoroaromatic structures, with this topic being less

explored for aliphatic organofluorines.

The physicochemical properties of these compounds are an important reason

behind their resistance to microbial catabolism and can be almost fully attributed

to the significant negative inductive effect associated with fluorine’s high

electronegativity. This creates a stereochemical and electronic unbalance on the

whole molecular structure, generally preventing the electrophilic attack of

molecular oxygen, which constitutes a primary step in most aerobic metabolic

pathways (Kiel and Engesser, 2015). Moreover, some fluoroorganics can act as

enzymatic inhibitors, being capable of irreversibly inhibiting enzymatic activity

(Neilson and Allard, 2002). Thus, their recalcitrance may also be due to their

capacity of inactivating their potential biocatalysts, preventing their

biotransformation. As a result of these characteristics, the biochemical interaction

between organofluorine compounds and microorganisms often results in their

incomplete degradation or no degradation at all (Neilson and Allard, 2002).

Complete defluorination of a fluoroorganic usually leads to its mineralization,

especially if occurring as a primary step on the catabolic pathway, since elimination

of fluoride is a critical step in the biodegradation of fluorinated compounds (Kiel

and Engesser, 2015). This is particularly relevant when fluorination occurs in core

structures of aromatic organofluorines or in short-chained aliphatic compounds, as

in both these cases fluorine’s inductive effects are more evident throughout the

whole molecular structure (Kiel and Engesser, 2015).

As defluorination capacity is a characteristic not commonly present in most

microorganisms, fluoroorganic compounds are more likely to be biodegraded

through unspecific reactions, as those observed in cometabolic pathways. Yet, even

in these conditions, fluorinated compounds might not be fully metabolized (Kiel

and Engesser, 2015). The concomitant presence of a fluorinated compound and a

co-substrate, might induce cometabolic reactions due to structure similarity or to

growth stimulation of the microbial population (Kiel and Engesser, 2015). The first

situation requires two substrates to be structurally related, and occurs when the

12

presence of the growth substrate is capable of inducing enzymes able to catalyse

the breakdown of its recalcitrant analogue, while the second situation corresponds

to the use of substrates capable of supporting microbial growth, leading to an

increase of catabolic enzymes and, consequently improving the chances of

degradation of the fluorinated substrate.

Aliphatic and aromatic organofluorines have distinct stereochemical and

biochemical demands when concerning their biological transformation, thus

exhibiting different pathways through which they may be metabolized. Aliphatic

fluoroorganics are generally smaller and chemically simpler than fluorinated

aromatics, with the exception of perfluorinated aliphatics that bear additional

functional groups or several ring structures that may influence their

biodegradability. As a result of their simpler structures, defluorinating reactions

are common primary steps in the microbial degradation of aliphatics, and some

different enzymes have been reported to catalyse such reactions (Fetzner and

Lingens, 1994). Concerning aromatic structures, it has been shown that their

biodegradation share some similarities with the degradation of their non-

fluorinated analogues, such as in the case of several phenols, benzenes, benzoates

and anilines, whose metabolic pathways are well established (Boersma et al., 2001;

Carvalho et al., 2006; Iwai et al., 2009). In these compounds, fluoride ion removal

is an essential step in their degradation because it facilitates the consequent

transformation of the resulting substrate and avoids the generation of unwanted,

dead-end metabolites, which may be more persistent or toxic than the parental

compound (Kiel and Engesser, 2015). Defluorination may occur before or after

fission of the aromatic ring, but may be hindered depending on the position where

the fluorine atom is on the aromatic ring or if there is more than one ring structure,

as in the case of polycyclic compounds (Neilson and Allard, 2002; Murphy et al.,

2009).

Genetic mechanisms are also an important factor in the microbial degradation and

transformation of fluoroorganic compounds. Gene transfer can, in some cases, lead

to the emergence of novel defluorinating pathways or endow non-metabolically

competent microorganisms with suitable catabolic mechanisms to attack

fluorinated molecules. The acquisition of such genotypes may be the result of

horizontal gene transfer or though the integration of functional replicons,

mediated by integrase enzymes (Janssen et al., 2001). Adaptation processes can

also be a way through which microorganisms acquire capacities to transform and

13

defluorinate fluoroorganic molecules, and gene transfer has an important influence

in such processes. Additionally, environmentally-driven genetic mutations, such as

recombinations, are also relevant in microbial enrichment and constitute important

adaptation processes to fluorinated xenobiotics (Janssen et al., 2001).

5. Genomic and metagenomic approaches in biodegradation studies

In biodegradation studies it is very important to properly identify the metabolically-

competent microorganisms as well as to understand their microbial dynamics.

Much of this valuable information is now more easily accessible, thanks to the the

development of omics tools. More specifically, genomic and metagenomic

approaches have allowed to deepen the investigation of the microbial world,

allowing a clearer identification of microbial species, and contributing to the

understanding of microbial community dynamics, also visualizing segments of the

microbiome (essentially uncultured microorganisms) which were invisible

otherwise.

Genomics corresponds to the study of gene function and structure, allowing

mapping and elucidating biological systems and reactions. In microbiology,

genomics revolutionized the taxonomy and phylogeny of microbial species through

the analysis of specific genes with taxonomic value. In bacteria, the 16S ribosomal

RNA gene (16S rRNA gene) has been widely used for the phylogenetic identification

of bacterial isolates as it is highly conserved within bacterial species, showing only

some variable regions (Coenye and Vandamme, 2003). Prior to 16S rRNA gene

sequence analysis, bacterial taxonomy was based on morphological, biochemical

and physiological characteristics of microbial strains, which was often a subjective

procedure (Handelsman, 2004). In biodegradation/bioremediation studies, 16S

rRNA gene sequencing allowed to more accurately identify bacterial species with

biodegradation capacities. A myriad of microorganisms capable of remediating and

neutralizing numerous environmental contaminants have been identified thanks to

this genomic-based approach.

Metagenomics is the application of genomic-based principles for the analysis of

microbial communities directly derived from environmental samples. One major

advantage of this tool, when compared with other genomic approaches, is that it

enables the combined analysis of cultured and non-cultured microorganisms,

14

allowing understanding the microbial composition within a whole community.

Through this approach, it is possible to obtain high-resolution genetic information

of complex microbial systems, such as community shifts and dynamics and

microbial composition, diversity and structure (Bell et al., 2013). Therefore, this

type of approach allows better understanding how a microbial community responds

and adapts to the presence of a target contaminant (Bell et al., 2013). For example,

microbial diversity has been regarded as a good indicator of ecosystem function

with environmental microbiomes with high levels of microbial diversity being

usually more resistant to anthropogenic disturbances (Bissett et al., 2007; Allison

and Martiny, 2008). Metagenomic approaches allow understanding how

environmental microbiomes are affected by the presence of contaminants, which

is very important for the assessment of the environmental impact caused by these

compounds.

6. Aim and outline of this thesis

The utilization of fluoroorganic compounds is increasing worldwide, accompanied

by a proportional increment on their environmental presence and distribution. Due

to the fact that the majority of these compounds are emergent pollutants, a lot is

yet to be known regarding their biodegradability and hazardous nature.

Consequently, knowledge on the biodegradability of these compounds and on

suitable bioremediation technologies capable of mitigating the environmental

impact of fluorinated xenobiotics is urgently needed. In this context, the work

developed in this master dissertation focused in the investigation of the

biodegradation of fluoroorganic compounds with different structures (aliphatics

and aromatics) and applications.

The present thesis is structured as follows: in Chapter 1, a general introduction is

provided, presenting the state of the art concerning the properties, applications

and biodegradation of fluoroorganic compounds and also outlining key concepts

and definitions associated with biodegradation and bioremediation of xenobiotics;

the experimental approach contemplated in this master thesis is presented in

Chapters 2 and 3. In Chapter 2, the biodegradation of structurally related aliphatic

carboxylic fluoroorganics with many industrial applications is explored, while in

Chapter 3, the biodegradation of an aromatic structure, a widely used veterinary

fluoroquinolone, when present individually and in mixture with a second antibiotic,

15

a veterinary cephalosporin, is investigated. In Chapter 4, some final remarks are

presented, including a general discussion on both experimental works and main

conclusions.

Both experimental works integrated in this thesis were submitted to international

peer-reviewed scientific journals, with the following references:

1. Alexandrino DAM, Mucha AP, Almeida CMR, Gao W, Jia Z and Carvalho MF.

(2016). Biodegradation of the veterinary antibiotics enrofloxacin and ceftiofur and

associated effects on microbial community dynamics. Submitted to SCIENCE OF THE

TOTAL ENVIRONMENT.

2. Alexandrino DAM, Ribeiro I, Pinto LM, Cambra R, Oliveira RS, Pereira F and

Carvalho MF. (2016). Biodegradation of mono-, di- and trifluoroacetate by microbial

cultures with diferente origins. Submitted to NEW BIOTECHNOLOGY.

16

2 CHAPTER BIODEGRADATION OF MONO-, DI- AND TRIFLUOROACETATE

BY MICROBIAL INOCULA WITH DIFFERENT ORIGINS

(submitted to NEW BIOTECHNOLOGY)

17

1. Introduction

Due to the useful properties that fluorine confers to organic molecules, the use of

synthetic organofluorines for industrial, medical and agricultural applications has

been significantly increasing in the last decades (Kiel and Engesser, 2015). As a

result of their vast applications, fluoroorganic molecules are becoming pollutants

of several environmental compartments, where they may persist for long periods

of time due to the recalcitrant nature of many of these molecules (Banks et al.,

1994; Thayer, 2006). The degradation of organofluorine compounds constitutes a

challenge to microorganisms not only because the environmental pollution

originated by these compounds is a relatively recent problem, causing

microorganisms to be exposed to compounds so far unknown, but also because

the C-F bond of organofluorines has one of the highest known energies, making it

challenging to cleave (O'Hagan, 2008).

Fluoroacetates (FAs) are a family of carboxylic aliphatic organofluorines composed

by mono- (MFA), di- (DFA) and trifluoroacetate (TFA) that are highly soluble in water,

non-volatile and, as a result, likely to be mobile in the environment. MFA is a

naturally-occurring organofluorine and its synthetic form is used in some countries

as a vertebrate pesticide. This compound is highly toxic, especially to mammals,

where it acts as a potent inhibitor of the tricarboxylic acid cycle (O'Halloran et al.,

2005; Camboim et al., 2012). A number of tropical and sub-tropical plants are

capable of producing and accumulating MFA, using it as a defence mechanism

against herbivores (Marais, 1944; O'Hagan et al., 1993; Davis et al., 2012), and a

few Streptomyces species have also been found to produce it (Sanada et al., 1986;

Deng et al., 2014). MFA is also an important building block and an intermediary

reagent used in the industrial synthesis of several fluorinated antibiotics and

synthetic aminoacids and is a secondary product resultant from the microbial

metabolism of several fluorinated pharmaceuticals and industrial reagents (Ihara

et al., 1996; Percy, 1997; Goncharov et al., 2006). DFA is used in the chemical

synthesis of various fluorinated compounds and is produced during the microbial

metabolism of a range of organofluorines (Fox et al., 1990; Visscher et al., 1994;

Ihara et al., 1996; Percy, 1997; Morii et al., 2004; Ge et al., 2007). This compound

is suggested to result from the thermolysis of several commercial fluorinated

polymers (Ellis et al., 2001). TFA is an important derivative of the tropospheric

degradation of several HCFCs and HFCs, and is also a resulting product from the

abiotic breakdown of fluorinated polymers (Martin et al., 2000; Ellis et al., 2002).

18

In addition, this compound is widely used as a building block for the production of

various synthetic fluoroorganic compounds (Tamura et al., 1993; Linderman et al.,

1994; Boivin et al., 1995).

FAs have been reported to occur in several environmental compartments, being the

aquatic media their major environmental sink (Wang et al., 2004). TFA has been

detected in seasonal wetlands, marine environments, rainwater and lotic

environments, in concentrations ranging from 30 to 600 ng L-1

(Cahill and Seiber,

2000; Cahill et al., 2001; Römpp et al., 2001; Frank et al., 2002; Scott et al., 2005).

Although current environmental concentrations of TFA appear to be non-toxic to

microorganisms and animals, presenting only mild toxicity to some plants, its

recalcitrance may eventually lead to the accumulation of higher concentrations,

thus increasing the potential for ecosystem damage (Berends et al., 1999; Bott and

Standley, 1999; Smit et al., 2009). The environmental occurrence of MFA is mainly

linked with its use as a pesticide, that is applied aerially or in baits, though releases

through discharges of chemical industries may also occur (Ogilvie et al., 2010). The

physicochemical properties of MFA (water solubility, lack of volatility and low Kow)

suggest considerable mobility in the environment, being likely to reach

groundwater streams and even surface waters. The environmental dynamics of DFA

remain poorly explored in the literature but its structural similarity to the other

FAs, namely regarding its physicochemical properties, suggests a similar

environmental behaviour.

MFA was found to be biodegraded by different soil microorganisms (Gentle and

Cother, 2014). Kelly (1965) reported for the first time the bacterial degradation of

MFA, and other MFA-degrading bacteria have been isolated afterwards (Meyer et

al., 1990; Emptage et al., 1997; Davis et al., 2012). Microbial degradation of this

compound is usually mediated by the enzyme fluoroacetate dehalogenase, which

catalyses the cleavage of the C-F bond in the molecule, yielding glycolate (Goldman,

1965; Kawasaki et al., 1992; Kurihara et al., 2000). Biodegradation of TFA has been

reported to occur under anaerobic conditions, though its aerobic conversion to

fluoroform has also been described (Visscher et al., 1994; Kim et al., 2000).

However, current results on TFA biodegradation lack reproducibility and, thus,

more studies are needed. DFA has been identified as a secondary metabolite

resultant from the anaerobic biodegradation of TFA, being further converted into

MFA and then acetate (Visscher et al., 1994). Though this data suggests

degradation of DFA under anaerobic conditions, to the best of our knowledge no

19

studies on the aerobic biodegradation of this compound are available in the

literature. Moreover, as these compounds may occur simultaneously in the

environment, it is important to understand how the degradation of each compound

is affected by the presence of its analogues. In this context, our work aimed to

investigate the aerobic biodegradation of MFA, DFA and TFA as sole carbon sources

and in mixtures of two FAs. In addition, co-metabolic degradation of DFA and TFA

in the presence of their non-fluorinated analogue, acetate, was also studied.

Biodegradation was investigated using microbial inocula from different origins.

2. Materials and Methods

2.1. Microbial inocula

Sediment and rhizosphere samples of Phragmites australis (Cav.) Trin. ex Steud.

were collected from a site in Estarreja, Portugal with a long history of industrial

chemical contamination (Oliveira et al., 2001), and used as an environmental

source of microorganisms. An activated sludge consortium originated from a

municipal wastewater treatment plant (Gondomar, Porto) was also used as

inoculum for this study. This inoculum was obtained by centrifuging 40 mL of

activated sludge (5000 rpm for 15 min at 4 ºC), washing twice the resultant pellet

with a minimal salts medium (MM) and resuspending it in the same medium to one

tenth of its original volume.

2.2. Biodegradation experiments

Biodegradation experiments were performed in batch mode in 250 mL flasks with

70 mL of sterile MM. MM contained (per litre of ultra-pure water): Na2HPO4•2H2O

2.7 g, KH2PO4 1.4 g, (NH4)2SO4 0.5 g, MgSO4•7H2O 0.2 g and 10 mL of a trace

elements solution with the following composition, per litre: Na2EDTA•2H2O 12.0 g,

NaOH 2.0 g, MnSO4•4H2O 0.4 g, ZnSO4•7H2O 0.4 g, H2SO4 0.5 mL, Na2SO4 10.0 g,

Na2MoO4•2H2O 0.1 g, FeSO4•7H2O 2.0 g, CuSO4•5H2O 0.1 g and CaCl2 1.0 g. Flasks

were inoculated with 5 g of fresh sediment or rhizosphere samples and for the

activated slugge consortium, flasks were inoculated in order to have an initial

optical density (OD) at 600 nm of 0.1. Cultures were fed with FAs individually, in

mixtures of two FAs and, for DFA and TFA, in cometabolism with acetate. When fed

20

individually, FAs were supplemented at a concentration of 20 mg L-1

(0.20, 0.17

and 0.15 mM for MFA, DFA and TFA, respectively), while in the binary mixtures of

FAs, each compound was fed at the concentration of 10 mg L-1

(0.10, 0.085 and

0.074 mM for MFA, DFA and TFA, respectively). Cultures in cometabolism with

acetate were supplemented with DFA or TFA at the concentration of 5 mg L-1

(0.042

and 0.037 mM for DFA and TFA, respectively) and fed three times a week with 500

mg L-1

of sodium acetate. In the latter treatment, cultures were weekly transferred

to new sterilised flasks in order to ensure sufficient oxygen for the aerobic

degradation of the target compounds. Biodegradation of FAs was followed during

a three week period, after which half of the cultures were transferred to new flasks

containing the same proportion of MM and re-fed with the respective carbon

sources. Cultures were incubated under aerobic conditions, in a rotary shaker (130

rpm) at 25 ºC in the dark. Abiotic controls consisting in MM supplemented

individually with each of the FAs (5 mg L-1

) and incubated under the same conditions

were also included. Experiments were conducted in duplicate. FAs biodegradation

was followed by periodically measuring bacterial growth and fluoride ion release.

2.3. Bacterial characterization of MFA-degrading cultures

The bacterial composition of MFA degrading cultures was analysed by spreading

several tenfold dilutions of culture samples onto minimal salts agar plates

supplemented with MFA as sole carbon source and Plate-Count Agar (PCA). The

plates were incubated at 25 ºC until growth was detected. Bacterial composition

was analysed by visual inspection and morphologically distinct colonies were

purified by streaking the different colonies in new agar plates.

2.4. Biodegradation capacity of bacterial isolates obtained from MFA-

degrading cultures

The capacity of the different bacterial strains isolated from the MFA-degrading

cultures to degrade this compound in axenic cultures was investigated by

inoculating single strains into 30 mL sterile flasks, filled to two thirds of their

volume with MM and supplemented with MFA at 20 mg L-1

. The initial OD (600 nm)

of the cultures was 0.1. Flasks were incubated in a rotary shaker (130 rpm at 25

ºC), in the dark. Biodegradation was followed along a three week period by

monitoring bacterial growth and fluoride ion release.

21

A bacterial culture consisting of a mixture of all MFA-degrading isolates was also

created and used as inoculum for investigating its capacity to degrade DFA and

TFA, fed individually as sole carbon source (20 mg L-1

) and in cometabolism with

MFA (20 mg L-1

of MFA and 5 mg L-1

of DFA or TFA).

2.5. Identification of MFA-degrading isolates

All the isolates capable of degrading MFA as single strains were identified through

16S rRNA gene sequence analysis. DNA was extracted from colonies obtained from

minimal salts agar plates supplemented with MFA or PCA plates, following a

standard phenol-chloroform extraction method, as described elsewhere (Sambrook

et al., 1989). Briefly, bacterial colonies were transferred to 1.5 mL microtubes to

which STE buffer (100 mM NaCl, 1 mM EDTA, 10 mM Tris/HCl, pH 8.0), sodium

dodecyl sulphate (20%) and proteinase K (20 mg mL-1

) were added. The mixture was

incubated overnight at 56 ºC with gentle shaking. After the incubation period,

samples were transferred to Light Phase Lock Gel tubes to which

phenol:chloroform:isoamyl (25:24:1) and chloroform:isoamyl (24:1) alcohols were

sequentially added, after intercalated centrifugations (14000 rpm for 3 minutes) to

ensure the separation of the aqueous and organic phases. Finally, the obtained

DNA was concentrated through ethanol precipitation and the resulting pellets were

dried under sterile conditions at room temperature. DNA extracts were then

dissolved in 50 µL of sterilised water.

Extracted DNA was amplified by Polymerase Chain Reaction (PCR) using the

universal primers 27F and 1492R (Weisburg et al., 1991). PCR reaction mixture

contained 2 µM of the universal primers, a Multiplex PCR Master Mix (Qiagen,

Valencia, CA) and template DNA sample. Negative controls were included and

consisted on the same PCR reaction mixture in which DNA was replaced by DNase,

RNase and protease-free water (5 Prime). PCR amplification conditions included

initial denaturation at 95 ºC for 15 minutes, followed by 30 cycles at 94 ºC for 30

seconds, 48 ºC for 90 seconds (annealing step) and 72 ºC for 2 minutes, and a final

extension at 72 ºC for 10 minutes. Amplification products were separated by

electrophoresis in a 1.5% agarose gel containing SYBR® Safe (ThermoFisher

Scientific, Massachusetts, USA) at 150 V for 30 minutes. DNA fragments were

visualised under UV light in a BioRad Molecular Imager® Gel Doc™ XR+ with Image

Lab™ Software and those showing amplification bands with a suitable size

22

(~1500bp) were sent for sequencing at i3S – Instituto de Investigação e Inovação

em Saúde (Porto, Portugal).

2.6. Analytical methods

Fluoride release was analysed by potentiometry, through the measurement of the

concentration of fluoride ion in the supernatant of culture samples, using a

fluoride-selective electrode (Crison 9655 C, Crison Instruments, S.A., Barcelona,

Spain). Prior to sample analysis, a calibration curve was constructed using

standards of sodium fluoride (0.001 to 1 mM) prepared in MM. A total ionic

strength adjustment buffer (TISAB III) was supplemented to the samples and

standards in a 1:10 ratio.

Microbial growth was monitored through the measurement in a spectrophotometer

(Model V-1200, VWR International, LLC, Pennsylvania, USA) of the optical density

(OD) at 600 nm of culture samples.

3. Results

3.1. Biodegradation of FAs by the different microbial inocula

The microbial capacity to degrade three structurally related FAs, MFA, DFA and TFA,

as sole carbon sources, in mixtures of two FAs and in cometabolism with acetate,

was investigated using microbial inocula with distinct origins. Fluoride release was

used as a key biodegradation indicator, since the main obstacle to the microbial

degradation of these compounds lies in the presence of this atom in their molecular

structures.

When supplemented as a sole carbon source, only MFA was degraded by the tested

microbial inocula. Activated sludge consortium showed complete defluorination of

MFA since the beginning of the experiment, whereas the treatments inoculated with

sediment or rhizosphere samples revealed a gradual increase in MFA degradation

performance (Fig. 3). In these latter cultures, total defluorination was also achieved:

for cultures inoculated with sediment samples this was obtained when fed a second

time with MFA, while for rhizosphere cultures this was observed in the following

feeding period (Fig. 3). Total defluorination of MFA was maintained in further MFA

23

feedings for an additional period of 2 months. None of the tested microbial inocula

were capable of defluorinating DFA or TFA, either when supplemented as sole

carbon sources or in cometabolism with acetate.

Biodegradation of MFA in mixture with DFA or TFA was also investigated. A mixture

of DFA and TFA was not considered since no biodegradation had been obtained

when these compounds were supplemented individually. When MFA was

supplemented with DFA, only a small fraction of fluoride was detected in the culture

medium of the different tested microbial consortia (Fig. 4).

Figure 3. Biodegradation performances, based on fluoride release, of MFA supplemented as a sole

carbon source during a two months period. White bars represent rhizosphere inoculum, grey bars,

sediment inoculum and black bars, activated sludge consortia. Days 21, 42 and 63 correspond to the

end of the 1st

, 2nd

and 3rd

MFA feeding periods, respectively. The results represent the mean of

duplicates and error bars show standard deviation.

The low concentration of released fluoride in these cultures indicates that the

simultaneous presence of the two FAs not only did not stimulate the

biodegradation of DFA, but also produced a negative effect in the biodegradation

of MFA, as the obtained fluoride concentration was not proportional to the

complete defluorination of this compound (Fig. 4). In the cultures supplemented

with a mixture of MFA and TFA, the concentration of fluoride ion analysed in the

culture medium was higher than that obtained in the cultures fed with MFA and

DFA, and the extent of fluoride released suggests that MFA was fully degraded as

0

10

20

30

40

50

60

70

80

90

100

21 42 63

% o

f M

FA

d

eflu

orin

atio

n

Time (days)

24

it is in agreement with its stoichiometric defluorination. This result suggests that,

unlike DFA, the presence of TFA in the mixture does not interfere with MFA

biodegradation and that, similarly to what happened with DFA, the addition of MFA

does not stimulate biodegradation of TFA. In the cultures fed with MFA, both as

sole carbon source or in mixture with TFA, a slight OD increase was observed (data

not shown), though for cultures inoculated with sediment or rhizosphere samples

this parameter could not be analysed along the first three feeding periods due to

the interference of the inocula in this analysis.

Abiotic controls were also established and followed in parallel with the

biodegradation experiments, revealing no fluoride release in any of the flasks

under the tested experimental conditions.

3.2. Characterization of MFA-degrading bacterial consortia and

biodegradation capacity of the isolated strains

All the cultures degrading MFA (individually or in mixture with TFA) were analysed

in terms of their bacterial diversity. A total of 43 bacterial isolates were obtained

from the degrading cultures: 12 strains were recovered from activated sludge, 15

strains from cultures inoculated with rhizosphere samples and 16 strains from

cultures inoculated with sediment samples (Table 2). All these isolates were tested

individually for their capacity to degrade MFA when supplemented as a sole carbon

source, revealing that out of the 43 isolates recovered, 15 were capable of

completely defluorinating MFA (Table 2). The highest number of MFA-degrading

isolates was obtained from activated sludge consortia.

A mixed culture composed by all MFA-degrading isolates was also established and

tested for its capacity to degrade DFA and TFA as sole carbon sources and in

cometabolism with MFA. Based on fluoride release, no biodegradation of DFA and

TFA, fed individually, was observed with this consortium. When MFA was

supplemented as a co-metabolite, the results obtained were very similar to the ones

previously observed with the mixtures of two FAs, i.e., the concentration of fluoride

ion analysed in the culture medium when MFA was fed with TFA correlated with the

total defluorination of MFA, suggesting that this defluorination pattern is attributed

solely to the degradation of MFA, but when DFA was present in the mixture,

biodegradation of MFA was inhibited and only ca. 10% of this compound was

defluorinated (Fig. 5).

25

Figure 4. Defluorination performance of the tested microbial consortia when supplemented with

mixtures of FAs after two feeding periods. A – activated sludge consortia; B – rhizosphere inoculum;

C – sediment inoculum. Black bars show expected fluoride concentrations considering complete

defluorination of both FAs in the mixture, grey bars show expected fluoride concentration considering

total defluorination of only MFA (represented as the molarity of the stoichiometric release of the

fluoride anion) and white bars show the concentration of fluoride ion released to the culture medium.

Results represent the mean of duplicates and error bars are relative to standard deviation.

0,000

0,050

0,100

0,150

0,200

0,250

0,300

0,350

0,400

MFA + DFA MFA + TFA

Flu

orid

e released

(m

M)

A

0,000

0,050

0,100

0,150

0,200

0,250

0,300

0,350

0,400

MFA + DFA MFA + TFA

Flu

orid

e released

(m

M)

B

0,000

0,050

0,100

0,150

0,200

0,250

0,300

0,350

0,400

MFA + DFA MFA + TFA

Flu

orid

e released

(m

M)

C

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

.

26

These results suggest that the culture consisting of the mixture of all MFA-

degrading isolates was unable to metabolise DFA and TFA, being capable of

defluorinating MFA in the presence of TFA, but not in mixture with DFA.

3.3. Identification of the MFA-degrading bacterial isolates

Bacterial isolates capable of degrading MFA as sole carbon source were identified

through 16S rRNA gene sequence analysis. The isolates were identified as 9

distinct species, belonging to different genera, mainly assigned to the

Proteobacteria phylum (Table 2). Activated sludge comprised MFA-degrading

isolates belonging to 6 genera: Stenotrophomonas, Herbaspirillum, Delftia,

Pseudomonas, Comamonas and Achromobacter. The genus Pseudomonas, as well

as the species Comamonas testosteroni and Achromobacter anxifer were present

in both activated sludge and cultures inoculated with sediment samples. An isolate

belonging to the genus Chryseobacterium were also obtained from these latter

cultures. In the cultures inoculated with rhizosphere samples, isolates capable of

degrading MFA as single strains were found to belong to Variovorax, Arthrobacter

and Pseudomonas (Table 2).

4. Discussion

Aliphatic organofluorines represent a class of compounds usually regarded as

common environmental pollutants (Neilson and Allard, 2002). The critical step in

the biodegradation of these compounds is the removal of fluoride ion (Kiel and

Engesser, 2015). Complete defluorination of FAs was reported to yield easily

degradable compounds that may be readily dissipated from the environment and

have no potential for ecosystems damage, such as glycolate, a known secondary

product of the biodegradation of MFA, or acetate, which is thought to result from

the anaerobic biodegradation of TFA (Visscher et al., 1994; Kurihara et al., 2000).

27

Table 2. Microbial strains isolated from the different MFA-degrading consortia and taxonomic identification of the microbial isolates capable of degrading MFA

as sole carbon source

Inoculum

Carbon source

supplemented to

the medium

Number of

microbial

isolates

recovered

Number of

isolates with

capacity to

degrade MFA

Identification of MFA degrading

microorganisms GenBank accession numbers

Activated sludge

MFA 5 3

Comamonas testosteroni strain MFA1 KX400799

Stenotrophomonas maltophili strain MFA2 KX400881

Herbaspirillum frisingense strain MFA4 KX756676

MFA and TFA 7 3

Delftia acidovorans strain MFA5 KX400852

Pseudomonas putida strain MFA15 KX400880

Achromobacter anxifer strain MFA16 KX398363

Rhizosphere

MFA 8 1 Pseudomonas sp. strain MFA9 KX404994

MFA and TFA 7 2

Variovorax paradoxus strain MFA10 KX400967

Arthrobacter humicola strain MFA12 KX400776

Sediment

MFA 8 1 Chryseobacterium taeanense strain MFA25 KX400798

MFA and TFA 8 3

Achromobacter anxifer strain MFA31 KX400775

Pseudomonas sp. strain MFA32 KX756677

Comamonas testosteroni strain MFA35 KX400851

28

0

0,05

0,1

0,15

0,2

0,25

0,3

0,35

0,4

0,45

0,5

0,55

MFA DFA TFA MFA + DFA MFA + TFA

Flu

orid

e released

(m

M)

Figure 5. Biodegradation based on fluoride release of FAs supplemented as sole carbon sources

and in cometabolism with MFA, by a mix of the 13 MFA-degrading microbial isolates. Dotted lines

indicate theoretical fluoride concentration corresponding to complete defluorination of the tested

compounds (represented as the molarity of the stoichiometric release of the fluoride anion).

Results represent the mean of duplicates and error bars show standard deviation.

Complete defluorination of MFA supplemented as a sole carbon source was

achieved in the cultures inoculated with activated sludge, rhizosphere and

sediment samples. While activated sludge cultures readily defluorinated the

supplemented MFA, an acclimation period was necessary for the other two

microbial consortia to degrade the target compound. This may be due to activated

sludge cells being in a more active metabolic condition than the other two microbial

consortia, as these communities are typically subjected to high organic loads,

having an easy access to growth substrates, and to high selective pressures,

essentially due to the presence of a wide range of organic molecules in

wastewaters. Nonetheless, the fact that the other tested microbial inocula also

degraded MFA, indicates that microorganisms capable of metabolising this

compound were originally present in these microbial consortia, though they

needed an acclimation period in order to prevail in the communities. Due to the

fact that no fluoride release was obtained in abiotic controls containing MFA,

defluorination observed in the tested cultures can be solely attributed to the

biological action of microorganisms in these cultures. Biodegradation of MFA has

been reported before (Meyer et al., 1990; Wong et al., 1992; Camboim et al., 2012;

Davis et al., 2012; Gentle and Cother, 2014). Most of the described MFA-degrading

microorganisms originated from environments where MFA was known to be

TFA

MFA + TFA

MFA

MFA + DFA

DFA

.

.

.

.

.

.

.

.

.

.

29

present, such as soils in the neighbourhood of MFA-producing plants or soils

adjacent to baits impregnated with this compound (Meyer et al., 1990; Gentle and

Cother, 2014). However, isolation of MFA-degrading microorganisms has also been

reported from samples not contaminated with this compound, indicating that the

microbial capacity to metabolise MFA is widespread in the environment and among

microorganisms (Wong et al., 1992; Camboim et al., 2012; Davis et al., 2012). This

is in agreement with our results, as the obtained MFA-degrading isolates originated

from environmental samples where MFA is not expected to be present.

DFA and TFA were not defluorinated by any of the tested cultures along an

enrichment period of ca. 4 months. The absence of TFA defluorination under

aerobic conditions is in agreement with the results reported by other authors, while

DFA biodegradation has never been investigated to the best of our knowledge

(Visscher et al., 1994; Benesch et al., 2002). Visscher et al. (1994) reported the

accumulation of a dead-end metabolite, identified as fluoroform, resultant from the

aerobic biodegradation of TFA. This metabolite still holds the trifluoromethyl group

in its structure, and is more toxic than the parent compound. Benesch et al. (2002)

found no aerobic biodegradation of TFA along a three month period by microbial

communities from vernal pool soils. Contrastingly, complete defluorination of TFA

under anaerobic conditions has been reported by Visscher et al. (1994) and Kim et

al. (2000), with TFA (in concentrations ranging from 0.2 to 51 mg L-1

) being

reductively dehalogenated under methanogenic conditions to DFA, MFA and

acetate. Co-supplementation of the microbial cultures with acetate, a compound

structurally similar to FAs and a common microbial substrate, did not produce a

positive effect in defluorination of DFA or TFA. The co-feeding of substrates with

chemical structures similar to their halogenated counterparts may have a positive

effect in their biodegradation, through the induction of metabolic enzymes capable

of acting on their metabolism. For example, a Burkholderia sp. strain was capable

of metabolising a group of mono- and di-chlorophenols in the presence of phenol

as a growth-supporting substrate, though for highly substituted chlorophenols,

such as trichlorophenols and pentachlorophenol this strategy was inefficient (De

Los Cobos-Vasconcelos et al., 2006). This may be due to the alteration of molecular

steric and biochemical properties that are associated with increasing halogenation

of organic compounds, which may cause differences in enzyme recognition and,

consequently, substrate interaction. The results obtained in our study indicate that

the enzymatic mechanisms involved in the degradation of acetate are not efficient

30

in the biodegradation of DFA or TFA. On the other hand, the addition of acetate to

the cultures fed with DFA or TFA could also have benefited the biodegradation of

these compounds by stimulating microbial growth, as reported for other

organofluorines (Amorim et al., 2014; Carvalho et al., 2016) which was not verified

in this study.

In order to understand how biodegradation is affected when two FAs are

simultaneously fed, MFA was supplemented to microbial inocula in mixture with

DFA or TFA. MFA defluorination was found to be negatively affected by the presence

of DFA in the culture medium, while TFA did not seem to exert any effect in the

biodegradation of this compound. This negative influence in MFA defluorination

may be associated with an enzymatic inhibition, as MFA and DFA share greater

stereochemical similarities than MFA and TFA. This could allow DFA to bind to the

active site of the enzyme that metabolises MFA, preventing the binding of this

compound to the enzymatic system, thus blocking its action and inhibiting

defluorination. As the trifluoromethyl moiety of the TFA molecule has a higher

steric bulk than DFA, the MFA degrading enzyme may have a higher capacity to

discriminate between these two compounds, and so inhibition does not occur. To

the best of our knowledge, the inhibitory effect of DFA in the metabolism of MFA

had never been reported before. The results obtained with MFA fed in mixture with

DFA or TFA also suggest that the metabolic enzymes responsible for the

biodegradation of MFA are selective for this compound and, thus, not able to attack

DFA or TFA. This selective MFA catabolism has been reported before by Donnelly

and Murphy (2008). The authors isolated a fluoroacetate dehalogenase from

Pseudomonas fluorescens strain and found that the enzyme was highly selective

for MFA and not capable of metabolising DFA and TFA. This is a clear example of

the impact that the degree of fluorination may have in the microbial metabolism of

fluorinated compounds.

A total of 13 bacterial strains with the capacity to degrade MFA as sole carbon

source were isolated from the different MFA-degrading cultures. Taxonomic

identification of these strains revealed several microbial species not linked before

with the biodegradation of MFA. Some of these species belong to the Pseudomonas

genus, which, according to previous studies, is known to accommodate a number

of MFA-degrading strains (Goldman, 1965; Donnelly and Murphy, 2008). The

bacterial isolates identified as Comamonas testosteroni, Variovorax paradoxus and

Delftia acidovorans, all belonging to the Comamonadaceae family, were also

31

capable of degrading MFA as sole carbon source, being isolated from all MFA-

degrading cultures, independently of their environmental origin. D. acidovorans

(formerly Moraxella sp.) is the only microbial isolate obtained in this study that has

been demonstrated before to degrade MFA (Kawasaki et al., 1992; Sota et al., 2002;

Kurihara and Esaki, 2008). C. testosteroni and V. paradoxus have never been

associated with the biodegradation of MFA, but their capacity to degrade other

recalcitrant compounds, including several chlorinated aromatics, has been

described before (Sylvestre, 1995; Boon et al., 2000; Bathe et al., 2009; Satola et

al., 2013). On the other hand, H. frisingense has never been implicated in the

biodegradation of environmental contaminants, to the best of our knowledge.

According to the literature, defluorination of MFA is catalysed by fluoroacetate

dehalogenase (Goldman, 1965; Kawasaki et al., 1992; Kurihara et al., 2000). As the

genetic expression of this enzyme generally occurs at the plasmidic level, it is

possible that horizontal transfer of this genotype may have occurred in the MFA-

degrading bacterial communities, which may have contributed to the significant

number of bacterial strains capable of degrading this compound obtained in our

study (Kawasaki et al., 1981; Kawasaki et al., 1992; Sota et al., 2002; Kurihara and

Esaki, 2008). The combination of all MFA-degrading isolates proved to be

ineffective in the metabolism of DFA and TFA, namely concerning defluorination of

these compounds, reinforcing the conclusion that the enzyme responsible for the

defluorination of MFA is unable to act on its di or tri-fluorinated counterparts.

Overall, the results obtained in this study call the attention to the recalcitrant nature

of DFA and TFA, as well as to the potential deleterious effects that their continuous

release into the environment may have. Though literature studies show that TFA

causes no or slight toxic effects in the environment, its increasing environmental

release is expected to cause accumulation of this compound, especially in aqueous

resources, which may lead to unknown consequences. The effects of the

environmental accumulation of DFA are not yet known, but its resistance to

biodegradation together with the fact that it may interfere in the degradation

mechanisms of defluorinating enzymes, deserves further attention. The inhibition

of MFA defluorination caused by the addition of DFA, verified in our experiments,

must be taken into consideration regarding the biological removal of mixtures of

structurally related fluorinated compounds.

32

5. Conclusion

The work developed in this study showed that MFA can be metabolised by several

bacterial strains from different environmental sources, and that the mechanisms

responsible for its catabolism of do not apply in the biodegradation of its di and

tri-fluorinated counterparts. Most of the obtained MFA-degrading isolates have not

been linked before to the biodegradation of MFA, expanding the range of known

microbial species capable of metabolising this fluoroaliphatic compound. Under

aerobic conditions, DFA and TFA were recalcitrant to microbial degradation and co-

supplementation with the structurally related and more easily degradable

substrates, acetate and MFA, had no effect in their biodegradation. These results

indicate that the degree of fluorination of fluoroaliphatic compounds significantly

influences their biological degradation. When present in mixture, DFA inhibited

MFA defluorination, while TFA did not produce any negative effect, a result that, to

our knowledge, had never been reported. Such interactions should be taken into

account when considering the biodegradation of mixtures of structurally similar

fluorinated compounds. Overall, this work emphasizes the recalcitrant nature of

DFA and TFA and the potential negative interactions induced by mixtures of

fluoroorganics. The persistence and accumulation of FAs in the environment is a

relevant issue and may potentially lead to ecosystems disturbances.

33

3 CHAPTER BIODEGRADATION OF THE VETERINARY ANTIBIOTICS

ENROFLOXACIN AND CEFTIOFUR AND ASSOCIATED

MICROBIAL COMMUNITY DYNAMICS

(submitted to SCIENCE OF THE TOTAL ENVIROMENT)

34

1. Introduction

Veterinary drugs are commonly used to treat numerous animal diseases.

Antibiotics constitute one of the most representative groups of these

pharmaceuticals, being used not only for the treatment and prevention of diseases,

but also for the promotion of animal growth and improvement of the nutritional

value of animal-based foodstuffs, despite the legal restrictions concerning these

latter applications (Cromwell, 2002; Li et al., 2011).

The overuse of veterinary drugs has contributed to the emergence of these

products in several environmental compartments, essentially as a result of the

employment of contaminated livestock waste as natural fertilizers (Loke et al.,

2000; Tasho and Cho, 2016). In addition, these drugs are also released in the

environment through wastewater treatment plants (WWTPs) effluents, because

WWTPs are, in most cases, not capable of dealing with this type of contaminants,

resulting in incomplete or even no removal of these compounds from agro-

industrial effluents (Corcoran et al., 2010).

Pharmaceuticals may be released to the environment in their parental form or as

metabolites, including some biologically active ones, and, since they are designed

to induce specific physiological and biochemical effects on their target organisms,

the environmental presence of these compounds can cause a wide range of toxic

effects (Sarmah et al., 2006). For the particular case of antibiotics, their

environmental presence may also promote the selection of antibiotic-resistant

microorganisms (Martinez, 2009). Fluoroquinolones (FQ) and cephalosporins (CP)

are two of the most widely used antibacterial pharmaceuticals worldwide. In 2012,

the consumption in Europe of both FQ and CP accounted for over 20% of the total

antibiotics consumption (Weist et al., 2014). FQ are piperazinyl derivatives of the

N-heterocyclic antibacterial compounds designated as quinolones (Felczak et al.,

2014). Their mode of action relies on the ability to inhibit the activity of

topoisomerases type II and IV, key enzymes in DNA replication, which leads to the

blockage of microbial cell multiplication (Hu et al., 2007). CP are semi-synthetic

analogous of the naturally-produced cephalosporin-C (Rex and Susan, 2002). Being

a class of β-lactam antibiotics, their antibacterial activity resides in their capability

to disrupt peptidoglycan biosynthesis affecting bacterial-cell integrity (Rex and

Susan, 2002). Both classes of antibiotics have a broad-spectrum activity towards

several aerobic and anaerobic pathogens. FQ have been widely reported to occur

in both terrestrial and aquatic ecosystems in trace concentrations, typically ranging

35

from ng L-1

to µg L-1

, though concentrations of several mg L-1

have also been

reported (Picó and Andreu, 2006; Larsson et al., 2007; Zhang and Li, 2011).

Physicochemical properties of CP promote a faster environmental dissipation of

these antibiotics, leading to lower residence times of these pharmaceuticals in the

environment (Junker et al., 2006) and lower detections. As a consequence of the

environmental release of these two classes of antibiotics, an increasing number of

microorganisms resistant to these drugs has been reported in the literature

(Miranda and Castillo, 1998; Walsh, 2000; Ho et al., 2001; Hooper, 2002; Su et al.,

2008), highlighting the importance of studying their biodegradation potential.

In this context, the main objective of this work was to investigate the

biodegradation of two veterinary antibiotics representative of the FQ and CP

groups, enrofloxacin (ENR) and ceftiofur (CEF), respectively. ENR has been reported

to occur in wastewaters, agricultural soils and animal manure, while several

metabolites of CEF resultant from animal detoxification have been detected in

manure and soils (Rex and Susan, 2002; Zhao et al., 2010; Sim et al., 2011; Li et

al., 2014). Degradation of these compounds mainly focuses in physicochemical

processes (Sturini et al., 2012; He et al., 2014; Zamanpour and Mehrabani-

Zeinabad, 2014; Yang et al., 2016), while less studies are found in the literature

concerning their biodegradation (Martens et al., 1996; Wetzstein et al., 1997; Rafii

et al., 2009; Erickson et al., 2014). In the present work, biodegradation of ENR and

CEF, supplemented individually and in mixture, was investigated using microbial

communities from the rhizosphere of plants derived from experimental

constructed wetlands used for the treatment of livestock wastewaters

contaminated with these antibiotics (100 µg L-1

) (unpublished data). The effect of

the target antibiotics in the microbial dynamics of the degrading cultures was also

studied through metagenomics analysis.

2. Materials and methods

2.1. Enrichment of microbial degrading cultures

Microbial cultures capable of degrading ENR and CEF were obtained by selective

enrichment of inoculated culture medium with the target antibiotics, supplemented

either individually or in mixture, and using acetate as a co-substrate. Rhizosphere

sediment samples obtained from experimental constructed wetlands previously

36

designed for the treatment of livestock wastewaters contaminated with the target

antibiotics were used as inocula. Enrichments were conducted in duplicate, in batch

mode and under aerobic conditions, during ca. 5 months. For that, 250 mL flasks

containing 50 mL of sterile minimal salts medium (MM) were inoculated with 5 g of

sediment and fed with the target antibiotics at the concentration of 1 mg L-1

and

acetate at the concentration of 400 mg L-1

. MM contained (per liter): ):

Na2HPO4•2H2O 2.7 g, KH2PO4 1.4 g, (NH4)2SO4 0.5 g, MgSO4•7H2O 0.2 g and 10 mL

of a trace elements solution with the following composition, per litre:

Na2EDTA•2H2O 12.0 g, NaOH 2.0 g, MnSO4•4H2O 0.4 g, ZnSO4•7H2O 0.4 g, H2SO4

0.5 mL, Na2SO4 10.0 g, Na2MoO4•2H2O 0.1 g, FeSO4•7H2O 2.0 g, CuSO4•5H2O 0.1 g

and CaCl2 1.0 g. Microbial cultures were incubated in a rotary shaker (130 rpm), at

25ºC and protected from light. Acetate was fed to the cultures twice a week. Every

3 weeks, 25 mL of the microbial cultures were transferred to new flasks containing

equal volume of MM and re-fed with the target antibiotics and acetate. Every week,

cultures were transferred to new flasks to assure appropriate aerobic conditions.

Microbial enrichment was followed by monitoring microbial growth, fluoride ion

release for ENR and by measuring the concentration of ENR and CEF in the culture

medium.

2.2. Biodegradation of different concentrations of ENR and CEF

After the enrichment period, biodegradation of the target antibiotics was

investigated for concentrations of 2 and 3 mg L-1

. For that, 250 mL flasks containing

25 mL of MM and 25 mL of the microbial cultures enriched in the previous phase

(section 2.1) were initially supplemented, in triplicate, with the target antibiotics,

each at a concentration of 3 mg L-1

(supplemented individually and in mixture) and

acetate (supplemented twice a week at a concentration of 400 mg L-1

). Cultures

were incubated for a 3 weeks period in the same conditions used during the

enrichments (section 2.1). Aerobic conditions were maintained in the microbial

cultures as described previously. Biodegradation was monitored by analysing

microbial growth, fluoride ion release for ENR and antibiotics concentrations in the

culture medium. At the end of the 3 weeks period, microbial cultures were again

diluted to half of their volumes and doped a second time with the target antibiotics

at the same concentration (3 mg L-1

) and acetate (supplemented in the same

regime). Biodegradation was followed for an additional 3 weeks period, after which

37

the same procedure was repeated to test the biodegradation of the antibiotics at a

lower concentration, each at 2 mg L-1

.

In parallel with the biodegradation experiments, two sets of abiotic controls were

established. One consisted in sterile MM supplemented with ENR and CEF, both

individually and in mixture, at a concentration of 2 mg L-1

, and the other consisted

in sterile MM inoculated with autoclaved microbial consortia obtained from the

enrichment phase (initial optical density at 600 nm of 0.1), supplemented with 2

mg L-1

of the target antibiotics. Controls were established in triplicates and

incubated for one month in the same conditions of the degradation experiments.

2.3. Analytical methods

Biomass growth was monitored by reading the absorbance of culture samples at

600 nm, in a spectrophotometer (V-1200, VWR International, USA).

Fluoride ion release was measured as an indicator of ENR defluorination. The

concentration of fluoride ion in solution was analyzed, after centrifuging samples

at 13000 rpm for 15 min, with a fluoride-selective electrode (Crison 9655 C, Crison

Instruments, S.A., Spain). Prior to sample analysis a calibration curve was obtained

using standard solutions of sodium fluoride (0.001 to 1 mM) prepared in MM. A

total ionic strength adjustment buffer (TISAB III) was supplemented to the samples

and standards in a 1:10 ratio.

CEF and ENR were analyzed in the supernatant of the culture samples by HPLC.

Supernatants were obtained through centrifugation at 13000 rpm for 15 min.

Separation of the target antibiotics was performed in a C18 Luna column (150 x

4.6 mm) from Phenomenex, coupled to a Beckman Coulter HPLC equipped with a

diode array detector (module 128) and an automatic sampler (module 508).

Chromatographic conditions were the same as described elsewhere (Cavenati et al.,

2012). ENR was screened at 280 nm, while CEF was detected at 290 nm. The

analytical detection limit (LOD) for all the target antibiotics was 0.1 mg L-1

. Standard

solutions of the antibiotics were prepared in MM (0.1 - 6 mg L-1

) and used to obtain

calibration curves prior to every analysis.

38

2.4. Analysis of the structure of the microbial communities

The effect of the enrichment process with the target antibiotics in the different

degrading cultures was investigated by comparing the structure of the microbial

communities of the soil samples used as initial inocula with that of the microbial

cultures obtained at the end of the biodegradation experiments. DNA from the soil

samples used as inocula for the experiments was extracted from 0.5 g (wet weight)

of homogenized sediment using PowerSoil® DNA Isolation Kit from MOBIO

Laboratories, Inc., according to the manufacturer’s instructions. DNA from the

degrading cultures was obtained using a standard phenol-chloroform extraction

method, as described elsewhere (Sambrook et al., 1989). Briefly, microbial biomass

was harvested by centrifuging 1 mL culture aliquots and removing the supernatant,

to which it was added STE buffer (100 mM NaCl, 1 mM EDTA, 10 mM Tris/HCl pH

8.0), sodium dodecyl sulphate (20%) and proteinase K (20 mg mL-1

). The mixture

was incubated overnight at 56ºC with gently shaking. After the incubation period,

samples were transferred to Light Phase Lock Gel tubes (5 Prime Inc., Hamburg,

Germany) to which phenol: chloroform: isoamyl (25:24:1) and chloroform: isoamyl

(24:1) alcohols were sequentially added, with intercalated centrifugations (14 000

rpm for 3 min) to separate the aqueous and organic phases. Finally, the obtained

DNA was concentrated through ethanol precipitation and the resulting pellets were

air dried in sterile conditions, at room temperature. DNA extracts from soil samples

and from the degrading-microbial consortia were then dissolved in 50 µL of

sterilized water.

Structure of microbial communities in different samples was assessed by Illumina

Miseq sequencing of the 16S rRNA gene. Fusion primers consists of adaptor A or

B, key sequence, barcode and template specific sequences were used in this study.

Specifically, the V4-V5 region of the bacterial 16s rRNA gene was amplified by

Polymerase Chain Reaction (PCR) with the forward primer 515F (5’-

GTGCCAGCMGCCGCGG-3’) and the reverse primer 907R (5’-

CCGTCAATTCMTTTRAGTTT-3’), and a 12 bp adaptor sequence was attached to the

5’ end of 515F. The 50 µL PCR reaction mixture contained 1 x PCR buffer (Mg2+

plus), 0.2 mM of each deoxynucleoside triphosphate, 0.4 mM of each primer, 1.25

U of TaKaRa Taq HS polymerase (TaKaRa Biotech, Dalian, China) and 1 µL template

DNA. The PCR amplification program included initial denaturation at 94°C for 5

min, followed by 32 cycles at 94°C for 30 s, 55°C for 30 s, and 72°C for 45 s, and a

final extension at 72°C for 5 min. Amplified products were subjected to

39

electrophoresis using a 1.8% agarose gel. Amplicon bands with a suitable size (475

bp) were excised from the gel and purified with an agarose gel DNA purification kit

(TaKaRa Biotech, Dalian, China). All of the purified amplicons were then combined

in equimolar amounts and submitted to high-throughput sequencing on an Illumina

MiSeq pyrosequencer. The MiSeq sequencing data was analysed using the

Quantitative Insights into Microbial Ecology (QIIME) pipeline (http://qiime.org/).

Briefly, low quality sequences, which have lengths of <200 bp, an average quality

score of <25 and primer mismatches were trimmed and the barcodes were

determined to assign sequence reads to the proper samples. Then, the chimeras

were detected using the UPARSE algorithm based on a database of chimera-free

sequences. The sequences, which were assigned to a mitochondrial or chloroplast

origin were eliminated with the Metaxa software tool and the V4–V5 region was

extracted with the V-Xtractor software tool.

2.5. Statistical analysis

For the biodegradation experiments, replicates of samples were analyzed

independently and mean values and corresponding standard deviations were

calculated. For the metagenomics analysis, composite samples were used on all

experimental conditions.

Statistical analysis was performed using the software STATISTICA version 12

(StatSoft, Inc., 2013). For antibiotic removals and ENR defluorination, statistically

significant differences were evaluated through a parametric Student’s t-test, using

mean values and corresponding standard deviations of the replicates. Statistical

significance was assumed when the p-value was below or equal to 0.05.

Metagenomics profiles were analyzed using PRIMER 6 software package (v. 6.1.11)

(Clarke and Gorley, 2006). Bacterial richness and diversity index (Shannon Index)

were calculated based on the different number of OTUs and relative abundances of

the different OTUs. Normalization of the metagenomics profiles was performed

using the presence/absence pre-treatment function and, afterwards, a resemblance

matrix was created using the Bray-Curtis similarity method, from which a

hierarchical cluster was constructed using group average method. SIMPROF test

was used to detect differences among generated clusters.

40

3. Results

3.1. Biodegradation of ENR and CEF

To investigate the biodegradation of ENR and CEF, an enrichment period of 5

months was conducted using sediment samples obtained from experimental

constructed wetlands treating livestock wastewaters contaminated with these

antibiotics. Acetate was added to the cultures as a growth supporting substrate.

The purpose of this acclimation phase was to allow the adaption of the microbial

communities to each antibiotic.

During the first nine weeks of the enrichment phase, both microbial growth and

defluorination were not followed in the cultures due to the interference of the

sediment inocula in the analysis of these parameters. According to Table 3, nine

weeks after the beginning of the enrichments, biodegradation of ENR (based on

fluoride release) in the cultures fed individually with this compound and in mixture

with CEF was ca. 53 % and 65 %, respectively. In these microbial cultures, ENR was

gradually defluorinated along each feeding period of 3 weeks, with most of fluoride

being released in the last two weeks (Table 3). These results remained very similar

until the end of the enrichment phase (data not shown), and the complete ENR

defluorination was never achieved. During this phase, CEF was found to be always

completely removed from the culture media, while ENR removals ranged between

45 and 55 % when supplemented individually or in a mixture, respectively. Along

the enrichment phase, microbial cultures always had an increase on their microbial

densities (supported by the addition of acetate), showing a gradual increment over

time in their optical density (OD) (data not shown).

After the enrichment period, microbial cultures were tested for their capacity to

degrade ENR and CEF at the concentrations of 2 and 3 mg L-1

. Microbial cultures

were initially fed with the highest concentration, 3 mg L-1

, to test their robustness

to degrade the target antibiotics. In these conditions, defluorination of ENR sharply

decreased, being obtained values of ca. 4 and 3 % in the cultures fed with ENR and

with a mixture of ENR and CEF, respectively (Fig. 6). However, based on antibiotics

analysis in supernatant culture medium, these microbial cultures were able to

consume ca. 40 % of the supplemented ENR (Fig. 6).

41

Table 3. Defluorination performance along a feeding period of 21 days, obtained nine weeks

after the beginning of the enrichment phase, for ENR supplied individually and in mixture with

CEF, at the concentration of 1 mg L-1

Time (days)

% of ENR defluorination

ENRa

ENR + CEFa

7 18 ± 1 6 ± 3

14 24 ± 6 46 ± 5

21 53 ± 2 65 ± 3

Note: a

Results are expressed as the mean of duplicates ± standard deviation

Removal efficiencies of 100% were always observed for CEF, both in the cultures

supplemented individually with this antibiotic and in the cultures fed concomitantly

with ENR (data not shown).

When the cultures were fed with 2 mg L-1

of the target antibiotics, ENR

biodegradation performance improved, namely its defluorination, despite the

attained values being far below those obtained during the enrichment phase with

1 mg L-1

. Under these circumstances, similar (p>0.05) defluorination efficiencies of

ENR were achieved in the cultures fed with ENR and with a mixture of the two

antibiotics, with values of 22 and 16 % of defluorination being obtained,

respectively. At this concentration, ENR removals were fairly constant, showing no

significant differences (p>0.05) to the ones obtained when this antibiotic was fed

at 3 mg L-1

(Fig. 6). Removal efficiencies of 100 % were again observed for CEF,

showing no significant differences in function of its concentration or the

concomitant presence of ENR.

The increase in antibiotics concentrations did not affect microbial growth, being

achieved OD increments similar to the ones observed in the enrichment phase (data

not shown). Analysis of the antibiotics in the supernatant of the microbial cultures

supplemented with 2 or 3 mg L-1

of ENR (both individually and in mixture with CEF)

revealed the presence of two metabolites, though in concentrations below the LOQ,

identified as ciprofloxacin (CIP) and norfloxacin (NOR) by comparison with the

corresponding standard solutions.

42

Comparing total removal with abiotic controls (Fig. 7), it was observed that a

substantial amount of CEF was removed abiotically, having also a considerable

capacity to adsorb to microbial cells. After 30 days of incubation, ca. 39 % and 37

% of CEF was removed in the controls with no cells and in the controls containing

autoclaved consortia, respectively (Fig. 7). In contrast, ENR showed no removal or

defluorination in the controls without cells, having only a slight potential for cell

adsorption as evidenced by the ca. 6 % removal obtained in the controls with

autoclaved cultures (Fig. 7). The adsorption behaviour of both antibiotics did not

seem to be influenced by their simultaneous presence, as no significant differences

were observed in this condition (p>0.05) (Fig. 7).

0

10

20

30

40

50

60

70

80

90

100

ENR ENR + CEF

% o

f EN

R r

em

oval/d

eflu

orin

atio

nENR removal

ENR defluorination

3 mg L

-1

2 mg L

-1

3 mg L

-1

2 mg L

-1

Figure 6. Biodegradation of ENR, supplied individually and in a mixture with CEF for the concentrations

of 3 and 2 mg L-1

. Results are expressed as the mean of triplicates and error bars are relative to standard

deviation.

43

3.2. Analysis of microbial communities’ dynamics

To investigate the effect of the enrichment process with the target antibiotics in

the microbial communities used as inocula for the degrading experiments,

microbial compositions at the beginning and at the end of the experiments were

compared by metagenomics analysis.

Cluster analysis based on the Bray-Curtis similarity method showed that the

enriched communities are significantly different from the initial ones and that the

mode of antibiotics supplementation (individually or in mixture) did not influence

the structure of the enriched microbial community (Fig. 8). This trend is also

supported by the clear differences determined among the initial and the enriched

consortia, with the latter showing lower microbial diversity and abundance (Table

4).

Concerning microbial structure, five dominant phyla were found in the initial

communities: Firmicutes, Proteobacteria, Actinobacteria, Bacteroidetes and

Chloroflexi, accounting for over 80% of the structure of the communities of the

initial inocula (Fig. 8). Microbial enrichments with the target antibiotics caused a

0

10

20

30

40

50

60

70

80

90

100

ENR CEF

% o

f an

tib

io

tic rem

oval

Removal obtained from the biodegradation experiments

Abiotic removal

Removal by cell adsorption (ENR and CEF doped individually)

Removal by cell adsorption (ENR and CEF doped in a mixture)

Figure 7. Removals of ENR and CEF obtained in different experimental conditions, for the concentration

of 2 mg L-1

. Results are expressed as the mean of triplicates and error bars show standard deviation.

44

clear decrease in the abundance of microorganisms belonging to the phyla

Firmicutes and Actinobacteria, while the phyla Proteobacteria and Bacteroidetes

gained expression, representing between 80 to 90% of the entire microbial

communities of the final consortia (Fig. 8). Cultures enriched with ENR and with a

mixture of ENR and CEF also showed an increase in microorganisms belonging to

the phylum Spirochaetae (Fig. 8).

Table 5 shows the relative abundance of the most represented taxonomic groups

identified in the initial microbial communities and in the antibiotics enriched

microbial cultures. Enrichments with ENR and CEF supplied individually and in a

mixture, led to the selection of microorganisms belonging to the taxonomic groups

Rhizobiales, Betaproteobacteria and Comamonadaceae and to the loss of

Acidomicrobiales, Anaerolineaceae and Xanthomonadaceae (Table 5). The bacterial

genus Dysgonomonas showed an increased expression with the antibiotics

enrichment, while the genus Clostridium lost representation in all the final

microbial communities (Table 5). The Betaproteobacteria class was the most

representative group in the enriched microbial communities, with the highest

number of unidentified species (ranging from 33.7 to 36.5 %).

Despite the general shifts observed at the genus level, for all the enriched microbial

communities, metagenomics analysis showed that the mode of antibiotics

supplementation led to the selection of specific genera. For the cultures enriched

with ENR, the selection of the genera Flavobacterium (20.8 %) and Achromobacter

(8.4 %) was observed, while the genera Stenotrophomonas (12.8 %) and

Chryseobacterium (29.3 %) increased their expression in the microbial cultures

enriched with CEF and with a mixture of ENR and CEF, respectively (Table 5).

4. Discussion

There are several physicochemical processes capable of removing FQ and CP from

environmental matrices, but only a few biotic mechanisms have been described for

their degradation (Sturini et al., 2012; He et al., 2014; Karlesa et al., 2014; Yang et

al., 2016). The potential of environmental microorganisms to biodegrade these

antibiotics is yet to be properly elucidated and the work developed in this study

intends to shed some more light in this respect.

45

Table 4. Diversity and abundance indexes of the initial inocula and microbial communities enriched

with the target antibiotics

Microbial community Richnessa

Diversityb

ENRinitial 368 3.967

CEFinitial 377 4.401

ENR + CEFinitial 410 4.757

ENRfinal 143 2.290

CEFfinal 145 2.253

ENR + CEFfinal 121 1.994

Note: a

number of OTU; b

Shannon diversity index (H’).

Microbial acclimation constitutes an important process in the biodegradation of

environmental pollutants, including pharmaceutical compounds. Liao et al. (2016),

compared the biodegradation performances of CIP by non-acclimated and

acclimated microbial communities, and showed that this antibiotic was more

readily removed by acclimated consortia. The 5-months enrichment phase

conducted in this study certainly had an important role in the biodegradation

performance of the target antibiotics at the tested concentrations, allowing the

selection of microorganisms with higher potential to deal with these compounds.

This is supported by the observed shifts in the diversity and richness of the

microbial communities after a prolonged time of enrichment.

ENR was shown in this study to be metabolized by the enriched microbial consortia,

though complete defluorination and removal of this antibiotic has never been

achieved. Microbial defluorination of this antibiotic was significantly influenced by

its concentration, with defluorination being higher when the antibiotic was

supplemented at 1 mg L-1

and declining markedly with the increase of ENR

concentration. However, under these circumstances, the removal efficiency of ENR

did not change significantly, suggesting that fluoride release constitutes a limiting

step in the biodegradation of ENR.

In the cultures fed with ENR (both individually and in mixture with CEF), the

metabolites CIP and NOR were consistently detected, but it remained unclear if

their production was a consecutive event or if it corresponded to independent

metabolic pathways.

46

Group average

MIX_T35

CEF_T35

ENR_T35

MIX_sed

CEF_sed

ENR_sed

Sam

ple

s

100806040

Similarity

Transform: Presence/absence

Resemblance: S17 Bray Curtis similarity

ENRinitial

CEFinitial

ENR + CEFinitial

ENRfinal

CEFfinal

ENR + CEFfinal

40% 60% 80% 100% 0% 20% 40% 60% 80% 100%

Similarity Relative abundance

Figure 8. Cluster analysis based on Bray-Curtis similarity of metagenomics profiles of microbial communities and relative abundance of the different bacterial phyla at the

beginning and at the end of the biodegradation experiments. Dashed lines indicate samples that are similar (p>0.05) according to the SIMPROF test.

■ Firmicutes ■ Chloroflexi

■ Proteobacteria ■ Planctomycetes

■ Actinobacteria ■ Deinococcus-Thermus

■ Bacteroidetes ■ Spirochaetae

47

Table 5. Metagenomics profiles of the initial inocula and enriched consortia, showing the relative abundance of each taxonomic group in the communities (relative abundances

below 2 % were not considered)

0% 5% 10% 15% 20% 25% 30% 35% 40%

Phylum Class Order Family Genus ENRinitial CEFinitial ENR+CEFinitial ENRfinal CEFfinal ENR+CEFfinal

Actinobacteria Acidimicrobiia Acidimicrobiales

Bacteroidetes Bacteroidia Bacteroidales Porphyromonadaceae Dysgonomonas

Flavobacteriia Flavobacteriales Flavobacteriaceae

Flavobacterium

Chryseobacterium

Chloroflexi Anaerolineae Anaerolineales Anaerolineaceae

Deinococcus-Thermus Deinococci Deinococcales Trueperaceae Truepera

Firmicutes Clostridia Clostridiales Christensenellaceae

Clostridiaceae Clostridium

Peptostreptococcaceae

Erysipelotrichia Erysipelotrichales Erysipelotrichaceae Turicibacter

Proteobacteria Alphaproteobacteria Rhizobiales

Bradyrhizobiaceae Bosea

Brucellaceae

Phyllobacteriaceae Mesorhizobium

Xanthobacteraceae

Betaproteobacteria

Burkholderiales Alcaligenaceae Achromobacter

Comamonadaceae

Variovorax

Neisseriales Neisseriaceae

Gammaproteobacteria Xanthomonadales Xanthomonadaceae

Arenimonas

Stenotrophomonas

Spirochaetae Spirochaetes Spirochaetales Spirochaetaceae Spirochaeta

48

Nonetheless, the identification of these metabolites in the cultures

supplemented with ENR suggests that, at least, part of the molecule is not

immediately subjected to an initial defluorination step, as the identified

metabolites (NOR and CIP) also bear a fluorine atom in their structures. The

metabolite CIP has been reported before to be involved in the biodegradation of

ENR by fungal species, being produced by deethylation of the ENR piperazine

ring (Wetzstein et al., 2006). On the other hand, to the best of our knowledge,

NOR has never been reported before as an intermediary metabolite of ENR

biodegradation. Further biodegradation of these two fluorinated metabolites are

described to proceed via attack to the piperazine ring, with fluoride removal

occurring afterwards through a hydroxylation reaction (Amorim et al., 2013; Liao

et al., 2016). In this study, it is possible that biodegradation of ENR follows a

similar pathway, in which the following steps may occur: (i) initial conversion of

ENR into CIP and/or NOR; (ii) loss of the piperazine moiety in both CIP and NOR,

resulting in other metabolites still bearing fluorine in their structure; (iii)

defluorination of these fluorinated products by hydroxylation. This chain of

reactions is expected to generate smaller and simpler molecules, with less

antibacterial activity that may be more easily used as carbon sources by

environmental microorganisms (Wetzstein et al., 2009; Liao et al., 2016). While

defluorination of ENR may not be an immediate catabolic step, it may contribute

to the inactivation of its bactericidal properties, as it has been shown before for

other FQ (Carvalho et al., 2016).

Studies on the biodegradation of ENR indicate that this antibiotic is mainly

degraded by fungi. Gloeophyllum striatum was reported to metabolize 5 and 10

mg L-1

of ENR, but complete degradation has never been achieved in a period of

eight weeks, while Mucor ramannianus was able to degrade ca. 79 mg L-1

of ENR

in a 21 days period, though no information on defluorination of the molecule is

given in the study (Martens et al., 1996; Wetzstein et al., 1997; Parshikov et al.,

2000; Wetzstein et al., 2006). A wide network of metabolites resulting from the

different biodegradation pathways of ENR by G. striatum has been identified,

with a small portion of these metabolites being non-fluorinated congeners of the

parental compound that were generated as a primary metabolic step through a

hydroxylation reaction (Wetzstein et al., 1997; Karl et al., 2006). Parshikov et al.

(2000) identified three fluorinated ENR metabolites produced by M.

ramannianus, with one of them being implicated before in the biodegradation

49

of ENR by G. striatum and another being involved in the biodegradation of CIP

also by M. ramannianus (Wetzstein et al., 1997; Parshikov et al., 1999). Recent

studies on FQ biodegradation are mainly focused in second-generation FQ, such

as NOR, CIP and ofloxacin, though bacterial degradation of moxifloxacin, a

fourth generation FQ, has also been investigated (Girardi et al., 2011; Amorim

et al., 2013; Maia et al., 2014; Carvalho et al., 2016; Liao et al., 2016). These

studies indicate that the capacity to degrade FQ depends largely on the

microorganisms involved and the associated growth conditions. For example,

CIP has shown to be recalcitrant, along 93 days, in both aquatic and soil

ecosystems at a concentration of 20 mg L-1

, with only minimal degradation (0.9

%) being found in soil after that period of time (Girardi et al., 2011). However,

higher biodegradation performances for this same antibiotic have been verified,

when present in lower concentration ranges. Microbial communities isolated

from a biological activated carbon filter system designed for the treatment of

lake water contaminated with antibiotics were capable of growing in the

presence of 10 mg L-1

of CIP as a sole substrate (Liao et al., 2016). Also, an

Alphaproteobacteria strain, Labrys portucalensis strain F11, was able to convert

85% of 1 mg L-1

of CIP in 28 days (Amorim et al., 2013). This bacterial strain was

also capable of metabolizing other FQ, including ofloxacin, NOR and

moxifloxacin, in concentrations ranging from 1 to 10 mg L-1

(Amorim et al.,

2013; Maia et al., 2014; Carvalho et al., 2016). In these latter studies, and

similarly to our results, defluorination was also shown to be a limiting step in

the microbial degradation of the tested FQ.

Microbial cultures supplemented with CEF were always capable of completely

removing this compound from the culture medium, independently of its

concentration or the concomitant presence of ENR. Although a part of this

removal was due to abiotic processes, these results are in agreement with other

literature studies on the biodegradation of CEF. A wide group of anaerobic

bacterial strains obtained from bovine waste was shown to be able to fully

remove 5 mg L-1

of this antibiotic within 24 to 120 hours (Rafii et al., 2009).

Biodegradation of 10 mg L-1

of CEF by fecal microorganisms has also been

reported (Li et al., 2011; Erickson et al., 2014). Among these microorganisms, a

Bacillus cereus was capable of growing with concentrations of this antibiotic

above 100 mg L-1

(Erickson et al., 2014). Some of these microorganisms were

found to be capable of expressing β-lactamases, a group of enzymes that play a

50

fundamental role in the complete degradation of CEF (Rafii et al., 2009; Erickson

et al., 2014). It is possible that part of the removal of CEF obtained in this work

is a result of similar enzymatic activities, as the expression of β-lactamases in

environmental microorganisms is a very common phenotype (Rafii et al., 2009;

Bush and Jacoby, 2010; Erickson et al., 2014). It is reported that one of the

primary targets in CEF biodegradation is the β-lactam moiety, a mechanism that

may also have occurred in CEF degradation by the microbial consortia enriched

in this work (Li et al., 2011). This reaction may be responsible for a considerable

reduction of CEF antibacterial properties, as the antibiotic potential of CP rely

heavily on the integrity of their lactam ring (Rex and Susan, 2002).

In the microbial cultures supplemented simultaneously with ENR and CEF,

biodegradation performances of these compounds were very similar to the ones

obtained in the cultures fed individually with these antibiotics. This indicates

that the metabolic mechanisms responsible for CEF removal do not affect ENR

degradation and vice-versa, and that the enzymes responsible for the

metabolism of these two drugs are likely to be distinct. This result is highly

relevant, as it suggests that the concomitant environmental presence of these

two antibiotics will not hinder their microbial removal.

Both biotic and abiotic mechanisms played an important role in the removal of

CEF. This is also expected to occur in an environmental scenario, which might

explain why CP do not tend to persist in the environment. Two abiotic

mechanisms, namely hydrolysis and photolysis, have been reported to be

involved in the breakdown of CP, including CEF (Jiang et al., 2010; Li et al., 2011).

In this work, abiotic degradation of CEF might have occurred through a

hydrolysis mechanism, as the experiments were always conducted in the

absence of light. Unlike CEF, abiotic degradation of ENR was found to have a

minor role in the removal of this antibiotic, indicating that it was primarily

degraded through the catabolic action of the enriched microbial consortia. In

addition, abiotic controls with autoclaved consortia also showed that both ENR

and CEF tended to bind to microbial membranes, with CEF showing a higher

potential. While this may account as a removal mechanism, adsorbed antibiotics

may still have been metabolized in the degradation experiments, as adsorption

is usually a reversible reaction.

Enrichments with the target antibiotics, supplied individually or in mixture, had

a significant effect on the structure and diversity of the microbial communities.

51

Both individual and simultaneous presence of ENR and CEF is expected to

promote microbial selection in the communities, selecting those

microorganisms capable of breaking down these compounds. Diversity of all

enriched consortia decreased when compared with the corresponding initial

communities, which may be a consequence of microbial cultures being exposed

to higher antibiotic concentrations and to growth conditions different than those

of the experimental systems from where the inoculum samples were derived.

However, abundance was markedly higher in the enriched consortia, which may

be due to the frequent co-supplementation of microbial cultures with acetate as

an easily degradable carbon source, allowing a higher growth of the

communities selected by the presence of the target antibiotics. In a study

conducted by Fernandes et al. (2015) on the removal of the antibiotics ENR and

tetracycline in constructed wetlands microcosms, the authors found that the

presence of ENR (100 μg L−1

) did not induce significant long-term changes in

microbial abundance and diversity, but resulted in significant differences in the

microbial community structure. Liao et al. (2016) observed a decrease in

microbial abundance but similar diversity indexes (Shannon index) in the

presence of CIP. However CIP was supplemented as a sole carbon source, which

could explain the lower abundance in the communities, and microbial dynamics

was followed along a shorter period of time (28 days), which could have been

not enough to trigger significant diversity alterations in the microbial

communities. Girardi et al. (2011) has shown that longer exposure periods to

CIP (up to 65 days) can cause considerable community shifts.

Overall, two of the most dominant bacterial phyla present in the initial

communities, Firmicutes and Actinobacteria, suffered a considerable decrease

in the enriched consortia, with the phyla Proteobacteria and Bacteroidetes,

gaining a higher expression in the enriched communities. In a metagenomics

study conducted with CIP, microorganisms belonging to Proteobacteria and

Actinobacteria phyla were mainly selected, whereas Bacteroidetes and

Firmicutes species lost their expression (Liao et al., 2016). The fact that in both

studies a selection of microorganisms belonging to the phylum Proteobacteria,

was promoted, with a special emphasis on Betaproteobacteria, suggests that

members of this taxonomic group may have an important role in the

biodegradation of FQ. Among the phylum Bacteroidetes, representation of the

genus Dysgonomonas increased in the consortia enriched with the target

52

antibiotics, both individually and in mixture, indicating that this taxonomic

group likely has a role in the biodegradation of both ENR and CEF. Liao et al.

(2016) also found an increase of Dysgonomonas species in CIP-enriched

communities, suggesting that microorganisms belonging to this genus may be

involved in the biodegradation of FQ. Other bacterial genera selected in the

enriched communities were Flavobacterium, Chryseobacterium,

Achromobacter, Variovorax and Stenotrophomonas. These genera have already

been associated with the biodegradation of recalcitrant organic compounds,

many of them halogenated. For example, Achromobacter species have been

reported to be involved in the biodegradation of several sulfonamides (Li et al.,

2009; Xu et al., 2013; Reis et al., 2014); Flavobacterium species have been

reported to be capable of degrading the chlorinated pesticide,

pentachlorophenol (Hu et al., 1994; Lo et al., 1998); Variovorax species were

shown to metabolize several derivatives of phthalate and the pesticide linuron

(Prasad and Suresh, 2012; Horemans et al., 2013; Prasad and Suresh, 2015) and

Chryseobacterium and Stenotrophomonas species were described to be capable

of using a wide range of chlorinated and fluorinated pesticides, such as

flubendiamide, tetraclorophenol or DDT (Deng et al., 2015; Jadhav and David,

2016; Pan et al., 2016).

5. Conclusion

In this study, ENR and CEF were degraded at different extents by microbial

communities derived from experimental constructed wetlands designed to treat

wastewaters contaminated with trace amounts of the two antibiotics. While

complete removal of CEF was always achieved, ENR showed to be more

recalcitrant. Removal percentages for this latter antibiotic between 40 and 60 %

and defluorination percentages between 3 and 79 % were obtained, with

biodegradation being affected by the increase in its concentration. The

simultaneous supplementation of ENR and CEF did not affect the biodegradation

of these antibiotics. Contrarily to what was found for ENR, abiotic mechanisms

had a significant role in the removal of CEF, which may be one of the reasons

why this antibiotic has a faster dissipation in the environment. Microbial

dynamics associated to the enrichments with the target antibiotics revealed a

shift in the structure of the microbial communities, with a predominant selection

of microorganisms belonging to the phyla Proteobacteria (e.g., Achromobacter,

53

Variovorax and Stenotrophomonas genera) and Bacteroidetes (e.g.,

Dysgonomonas, Flavobacterium and Chryseobacterium genera). Overall, this

work demonstrated that microorganisms are capable of adapting and

responding to the presence of different emergent pollutants, like the antibiotics

used in this study, though concentration is a key factor in the biodegradation

process. The biodegradation capacity of the tested antibiotics exhibited by the

microbial communities enriched in this study suggest that environmental sites

contaminated with mixtures of ENR and CEF, where lower concentrations of

these contaminants are typically present, are likely to be recovered, at least

partially, through bioremediation processes.

4 CHAPTER GENERAL DISCUSSION AND CONCLUSIONS

55

1. General discussion

During the last decades, fluorinated organic compounds have become common

environmental contaminants due to their high versatility and favourable properties,

being increasingly used in various sectors of our societies. Nowadays,

fluoroorganics are amongst some of the most used synthetic compounds in areas

such as human and veterinary medicine, agriculture or even in the industrial sector.

While a lot of research and development is being carried out on the industrial

production of organofluorine compounds, less efforts are being directed towards

the effects of these products on human and environmental health. This is a

concerning issue, as the consumption of fluorinated products is not showing a

decreasing trend. Also, a lot of these products find themselves in legal grey areas,

as their production, application and elimination procedures remain highly

unregulated.

Only recently the hazardous nature of organofluorines, including their human

toxicity and potential for ecosystem damage, has been acknowledged in the

scientific literature. Key et al. (1997) was among the first scientists to address this

issue, being also the first author to recognize organofluorines as “ubiquitous

environmental contaminants”. Since then, other important works have been

published, but a big gap of knowledge still exists on the toxicity, environmental

impact and biodegradation of fluorinated compounds.

In the last 20 to 30 years, biodegradation studies mainly targeted non-halogenated

and chlorinated environmental contaminants, focusing less on organofluorine

compounds. Presently, due to the rapid expansion and spread of fluorinated

compounds, these are found in the environment as micropollutants, although

higher concentrations than those usually reported for emergent contaminants have

also been found (Larsson et al., 2007; Piekarz et al., 2007; Harada and Koizumi,

2009).

Knowledge on fluoroorganics biodegradation efficiency, metabolic pathways, as

well as on the involved microorganisms is highly relevant, as it is crucial for the

design of efficient bioremediation technologies. In this thesis, the biodegradation

of different organic fluorinated compounds was investigated, in order to

understand their biodegradation potential by environmental microbial

communities and to obtain knowledge on the microbial species/microbial

communities involved in the biodegradation process. Also, and whenever possible,

56

insights on the metabolic pathways of the target compounds were given, based on

the obtained experimental data.

The first study conducted showed that MFA was readily biodegraded by a wide

diversity of environmental bacteria. This capacity has been reported before by other

authors and may be due to the existence of an enzyme capable of specifically

catalysing the defluorination of this compound. Yet, such specific defluorinating

enzymes are not common in the metabolism of organofluorines, and in most cases

defluorination occurs as a result of non-specific catabolic reactions. Also, the

biodegradation results obtained for MFA, DFA and TFA, showed that the degree of

fluorination plays a major role in the recalcitrance of fluorinated compounds. For

the case of FAs, this resulted in MFA being completely defluorinated, while DFA

and TFA still held the fluorine atoms in their structures. The absence of a proper

aerobic biological degradation of DFA and TFA is concerning, as their recalcitrance

and environmental dynamics may lead to an increase of these compounds in

aquatic ecosystems, where they are likely to persist over time.

The work conducted with ENR revealed that biodegradation of this compound was

highly influenced by its concentration, with degradation efficiency decreasing with

the increase on the concentration of this compound. Additionally, results showed

that defluorination apparently is not a primary step in the biodegradation of this

fluoroquinolone, as the metabolic intermediates CIP and NOR, both still bearing

fluorine in their structures, were detected in the culture medium. While biological

mechanisms had a more preponderant role on ENR removal than on the removal of

CEF, biodegradation is expected to have an important role in the environmental

removal of both these antibiotics, even when present in a mixture. It was also

possible to attest that environmental microbial communities have the capacity to

adapt and respond to the presence of this type of contaminants, even in higher

concentrations than those usually reported for antibiotics.

An important aspect of the work developed in this thesis was the investigation of

the biodegradation of the target compounds when supplemented as mixtures of

xenobiotics. In natural environments, contaminants are usually present in complex

mixtures with other compounds, especially if they have similar sources of input

into the environment. Mixtures of xenobiotics are relevant, as they can have

increased deleterious effects in the environment due to synergetic interactions

between them. Also, from a microbial point of view, the presence of the target

compound in a complex mixture may alter metabolic dynamics, eventually

57

affecting overall biodegradation potential either due to metabolic inhibition or to

toxic effects induced in the degrading microorganisms. This was observed in the

study on the biodegradation of FAs, when MFA was fed in mixture with DFA, with

microbial defluorination of the first compound decreasing markedly in the

presence of the second, likely due to competitive substrate inhibition.

Acetate was used as a growth supporting substrate in the experimental work

conducted in this thesis, for two main reasons: (i) to investigate the biodegradation

of the target compounds in the presence of an easily accessible carbon and energy

source that could serve as a cometabolite and (ii) to mimic the organic carbon loads

usually present in some natural environments or in WWTPs. Cometabolism

constitutes an important mechanism in the biodegradation of recalcitrant

compounds, since the metabolic conversion of many of these compounds occurs

through fortuitous reactions promoted by the presence of highly energetic

substrates (Criddle, 1993). Also, in a real environmental scenario there is always

organic matter available for microbial consumption.

2. Conclusion

The experiments conducted under the scope of this thesis showed that, although

having increased resistance to biodegradation mechanisms when compared with

other xenobiotics, fluoroorganic compounds can be metabolized by environmental

microorganisms.

Several bacterial strains from distinct environmental sources were able to utilize

MFA as a sole carbon source, though DFA and TFA were shown to be recalcitrant

under different experimental conditions, indicating that the metabolic mechanisms

involved in the biodegradation of MFA are not able to act in the degradation of

these two compounds. The majority of these MFA-degrading bacterial strains have

never been linked before to the biodegradation of this compound, and so this work

shows for the first time the capacity of these microbial species to degrade this

fluoroaliphatic. It was also found in this work that DFA negatively affects MFA

microbial metabolism, which may be a limiting factor when considering the

biological recovery of environmental matrices contaminated with mixtures of these

compounds.

58

An enriched microbial consortium was capable of removing and defluorinating ENR

supplied in a range of concentrations between 1 and 3 mg L-1

, though at different

extents. Biodegradation of this compound markedly decreased with the increase in

its concentration and was not affected by the concomitant presence of CEF. On the

other hand, CEF biodegradation was not affected by the different concentrations

tested. This study also revealed that the microbial communities used as inocula

were capable of adapting and responding to the presence of these antibiotics. The

results obtained indicate that bioremediation of environmental sites contaminated

with mixtures of ENR and CEF may be possible, especially when assuming that

antibiotic concentrations lower than those tested in this study are typically present.

In overall, these studies emphasized the potential of environmental-occurring

microorganisms to biodegrade organofluorinated contaminants. Two main factors

were identified as crucial in the biodegradability of the tested fluorinated

compounds: the degree of fluorination and compound concentration. Microbial

cultures used in the two conducted studies have potential to be used in

bioremediation strategies of fluoroorganic compounds.

5 CHAPTER REFERENCES

60

1. Adams, G.O., Fufeyin, P.T., Okoro, S.E., and Ehinomen, I. (2015). Bioremediation,

Biostimulation and Bioaugmention: A Review. International Journal of Environmental

Bioremediation & Biodegradation 3(1), 28-39.

2. Allison, S.D., and Martiny, J.B.H. (2008). Resistance, resilience, and redundancy in microbial

communities. Proceedings of the National Academy of Sciences 105(Supplement 1), 11512-

11519.

3. Amorim, C.L., Ferreira, A.C.S., Carvalho, M.F., Afonso, C.M.M., and Castro, P.M.L. (2014).

Mineralization of 4-fluorocinnamic acid by a Rhodococcus strain. Applied Microbiology and

Biotechnology 98(4), 1893-1905.

4. Amorim, C.L., Moreira, I.S., Maia, A.S., Tiritan, M.E., and Castro, P.M.L. (2013).

Biodegradation of ofloxacin, norfloxacin, and ciprofloxacin as single and mixed substrates

by Labrys portucalensis F11. Applied Microbiology and Biotechnology 98(7), 3181-3190.

5. Banks, R.E., Smart, B.E., and Tatlow, J.C. (1994). Organofluorine chemistry: principles and

commercial applications. New York [etc.]: Plenum Press.

6. Bathe, S., Schwarzenbeck, N., and Hausner, M. (2009). Bioaugmentation of activated sludge

towards 3-chloroaniline removal with a mixed bacterial population carrying a degradative

plasmid. Bioresource Technology 100(12), 2902-2909.

7. Bégué, J.-P., and Bonnet-Delpon, D. (2006). Recent advances (1995–2005) in fluorinated

pharmaceuticals based on natural products. Journal of Fluorine Chemistry 127(8), 992-

1012.

8. Bell, T.H., Greer, C.W., and Yergeau, E. (2013). "Metagenomics Potential for Bioremediation,"

in Encyclopedia of Metagenomics, ed. E.K. Nelson. (New York, NY: Springer New York), 1-

11.

9. Benesch, J.A., Gustin, M.S., Cramer, G.R., and Cahill, T.M. (2002). Investigation of effects of

trifluoroacetate on vernal pool ecosystems. Environmental Toxicology and Chemistry 21(3),

640-647.

10. Berends, A.G., Boutonnet, J.C., Rooij, C.G.D., and Thompson, R.S. (1999). Toxicity of

trifluoroacetate to aquatic organisms. Environmental Toxicology and Chemistry 18(5),

1053-1059.

11. Bissett, A., Burke, C., Cook, P.L.M., and Bowman, J.P. (2007). Bacterial community shifts in

organically perturbed sediments. Environmental Microbiology 9(1), 46-60.

12. Boersma, M.G., Solyanikova, I., Van Berkel, W.J., Vervoort, J., Golovleva, L., and Rietjens, I.M.

(2001). 19F NMR metabolimics for the elucidation of microbial degradation pathways of

fluorophenols. Journal of Industrial Microbiology & Biotechnology 26(1/2), 22-34.

13. Boivin, J., El Kaim, L., and Zard, S.Z. (1995). Trifluoromethyl ketones from carboxylic acids.

Part II. A versatile access to trifluoromethylated heterocycles. Tetrahedron 51(9), 2585-

2592.

14. Boon, N., Goris, J., De Vos, P., Verstraete, W., and Top, E.M. (2000). Bioaugmentation of

Activated Sludge by an Indigenous 3-Chloroaniline-Degrading Comamonas testosteroni

Strain, I2gfp. Applied and Environmental Microbiology 66(7), 2906-2913.

15. Bott, T.L., and Standley, L.J. (1999). Incorporation of trifluoroacetate, a hydrofluorocarbon

decomposition byproduct, by freshwater benthic microbial communities. Water Research

33(6), 1538-1544.

16. Bush, K., and Jacoby, G.A. (2010). Updated Functional Classification of β-Lactamases.

Antimicrobial Agents and Chemotherapy 54(3), 969-976.

17. Cahill, T.M., and Seiber, J.N. (2000). Regional distribution of trifluoroacetate in surface

waters downwind of urban areas in Northern California. ACS Division of Environmental

Chemistry, Preprints 40(1), 5-7.

18. Cahill, T.M., Thomas, C.M., Schwarzbach, S.E., and Seiber, J.N. (2001). Accumulation of

Trifluoroacetate in Seasonal Wetlands in California. Environmental Science & Technology

35(5), 820-825.

19. Camboim, E.K.A., Almeida, A.P., Tadra-Sfeir, M.Z., Junior, F., #xed, G., c., et al. (2012).

Isolation and Identification of Sodium Fluoroacetate Degrading Bacteria from Caprine

Rumen in Brazil. The Scientific World Journal 2012, 6.

20. Carvalho, M.F., Ferreira, M.I.M., Moreira, I.S., Castro, P.M.L., and Janssen, D.B. (2006).

Degradation of Fluorobenzene by Rhizobiales Strain F11 via ortho Cleavage of 4-

Fluorocatechol and Catechol. Applied and Environmental Microbiology 72(11), 7413-7417.

61

21. Carvalho, M.F., Maia, A.S., Tiritan, M.E., and Castro, P.M.L. (2016). Bacterial degradation of

moxifloxacin in the presence of acetate as a bulk substrate. Journal of Environmental

Management 168, 219-228.

22. Cavenati, S., Carvalho, P.N., Almeida, C.M.R., Basto, M.C.P., and Vasconcelos, M.T.S.D.

(2012). Simultaneous determination of several veterinary pharmaceuticals in effluents from

urban, livestock and slaughterhouse wastewater treatment plants using a simple

chromatographic method. Water Science and Technology 66(3), 603-611.

23. Chang, R. (2010). Chemistry (10th

Ed., English). New York: McGraw-Hill.

24. Clarke, K.R., and Gorley, R.N. (2006). PRIMER v6: User Manual/Tutorial. PRIMER-E, Plymouth,

192.

25. Coenye, T., and Vandamme, P. (2003). Intragenomic heterogeneity between multiple 16S

ribosomal RNA operons in sequenced bacterial genomes. FEMS Microbiology Letters 228(1),

45-49.

26. Corcoran, J., Winter, M.J., and Tyler, C.R. (2010). Pharmaceuticals in the aquatic

environment: A critical review of the evidence for health effects in fish. Critical Reviews in

Toxicology 40(4), 287-304.

27. Crawford, R.L. (1998). Bioremediation: Principles and Applications. Cambridge University

Press.

28. Criddle, C.S. (1993). The kinetics of cometabolism. Biotechnology and Bioengineering

41(11), 1048-1056.

29. Cromwell, G.L. (2002). Why and how antibiotics are used in swine production. Animal

Biotechnology 13(1), 7-27.

30. Das, S., and Dash, H.R. (2014). "1 - Microbial Bioremediation: A Potential Tool for

Restoration of Contaminated Areas," in Microbial Biodegradation and Bioremediation.

(Oxford: Elsevier), 1-21.

31. Davis, C.K., Webb, R.I., Sly, L.I., Denman, S.E., and McSweeney, C.S. (2012). Isolation and

survey of novel fluoroacetate-degrading bacteria belonging to the phylum Synergistetes.

FEMS Microbiology Ecology 80(3), 671-684.

32. De Los Cobos-Vasconcelos, D., Santoyo-Tepole, F., Juárez-Ramírez, C., Ruiz-Ordaz, N., and

Galíndez-Mayer, C.J.J. (2006). Cometabolic degradation of chlorophenols by a strain of

Burkholderia in fed-batch culture. Enzyme and Microbial Technology 40(1), 57-60.

33. Dean-Ross, D., Moody, J., and Cerniglia, C.E. (2002). Utilization of mixtures of polycyclic

aromatic hydrocarbons by bacteria isolated from contaminated sediment. FEMS

Microbiology Ecology 41(1), 1-7.

34. Deng, H., Ma, L., Bandaranayaka, N., Qin, Z., Mann, G., Kyeremeh, K., et al. (2014).

Identification of Fluorinases from Streptomyces sp MA37, Norcardia brasiliensis, and

Actinoplanes sp N902-109 by Genome Mining. ChemBioChem 15(3), 364-368.

35. Deng, S., Chen, Y., Wang, D., Shi, T., Wu, X., Ma, X., et al. (2015). Rapid biodegradation of

organophosphorus pesticides by Stenotrophomonas sp. G1. Journal of Hazardous Materials

297, 17-24.

36. Donnelly, C., and Murphy, C.D. (2008). Purification and properties of fluoroacetate

dehalogenase from Pseudomonas fluorescens DSM 8341. Biotechnology Letters 31(2), 245-

250.

37. Ellis, D.A., Hanson, M.L., Sibley, P.K., Shahid, T., Fineberg, N.A., Solomon, K.R., et al. (2001).

The fate and persistence of trifluoroacetic and chloroacetic acids in pond waters.

Chemosphere 42(3), 309-318.

38. Ellis, D.A., Moody, C.A., and Mabury, S.A. (2002). "Trifluoroacetic Acid and Longer Chain

Perfluoro Acids-Sources and Analysis," in Organofluorines, ed. A.H. Neilson. (Berlin,

Heidelberg: Springer Berlin Heidelberg), 103-120.

39. Emptage, M., Tabinowski, J., and Odom, J.M. (1997). Effect of Fluoroacetates on

Methanogenesis in Samples from Selected Methanogenic Environments. Environmental

Science & Technology 31(3), 732-734.

40. Erickson, B.D., Elkins, C.A., Mullis, L.B., Heinze, T.M., Wagner, R.D., and Cerniglia, C.E.

(2014). A metallo-β-lactamase is responsible for the degradation of ceftiofur by the bovine

intestinal bacterium Bacillus cereus P41. Veterinary Microbiology 172(3–4), 499-504.

62

41. Felczak, A., Zawadzka, K., and Lisowska, K. (2014). Efficient biodegradation of quinolone –

Factors determining the process. International Biodeterioration & Biodegradation 96, 127-

134.

42. Fernandes, J.P., Almeida, C.M.R., Pereira, A.C., Ribeiro, I.L., Reis, I., Carvalho, P., et al.

(2015). Microbial community dynamics associated with veterinary antibiotics removal in

constructed wetlands microcosms. Bioresource Technology 182, 26-33.

43. Fetzner, S., and Lingens, F. (1994). Bacterial dehalogenases: biochemistry, genetics, and

biotechnological applications. Microbiological Reviews 58(4), 641-685.

44. Fox, B.G., Borneman, J.G., Wackett, L.P., and Lipscomb, J.D. (1990). Haloalkene oxidation by

the soluble methane monooxygenase from Methylosinus trichosporium OB3b: mechanistic

and environmental implications. Biochemistry 29(27), 6419-6427.

45. Frank, H., Christoph, E.H., Holm-Hansen, O., and Bullister, J.L. (2002). Trifluoroacetate in

Ocean Waters. Environmental Science & Technology 36(1), 12-15.

46. Ge, F., Wang, Z., Wan, W., Lu, W., and Hao, J. (2007). One-pot synthesis of 2-trifluoromethyl

and 2-difluoromethyl substituted benzo-1,3-diazoles. Tetrahedron Letters 48(18), 3251-

3254.

47. Gentle, M., and Cother, E. (2014). Biodegradation of 1080: Testing soils in south-eastern

Australia for sodium fluoroacetate-degrading micro-organisms. Ecological Management &

Restoration 15(1), 52-57.

48. Gilchrist, A. (2015). 10 Best-Selling Brand-Name Drugs in 2015 [Online]. Pharmacy Times.

Available: http://www.pharmacytimes.com/news/10-best-selling-brand-name-drugs-in-2015

[Accessed 18/07/2016].

49. Girardi, C., Greve, J., Lamshöft, M., Fetzer, I., Miltner, A., Schäffer, A., et al. (2011).

Biodegradation of ciprofloxacin in water and soil and its effects on the microbial

communities. Journal of Hazardous Materials 198, 22-30.

50. Goldman, P. (1965). The Enzymatic Cleavage of the Carbon-Fluorine Bond in Fluoroacetate.

Journal of Biological Chemistry 240(8), 3434-3438.

51. Goncharov, N.V., Jenkins, R.O., and Radilov, A.S. (2006). Toxicology of fluoroacetate: a

review, with possible directions for therapy research. Journal of Applied Toxicology 26(2),

148-161.

52. Gribble, G.W. (2003). The diversity of naturally produced organohalogens. Chemosphere

52(2), 289-297.

53. Handelsman, J. (2004). Metagenomics: Application of Genomics to Uncultured

Microorganisms. Microbiology and Molecular Biology Reviews 68(4), 669-685.

54. Harada, K.H., and Koizumi, A. (2009). Environmental and biological monitoring of persistent

fluorinated compounds in Japan and their toxicities. Environmental Health and Preventive

Medicine 14(1), 7-19.

55. Harnisch, J., and Eisenhauer, A. (1998). Natural CF4 and SF6 on Earth. Geophysical Research

Letters 25(13), 2401-2404.

56. Haynes, W.M. (2014). CRC handbook of chemistry and physics. CRC press.

57. He, X., Mezyk, S.P., Michael, I., Fatta-Kassinos, D., and Dionysiou, D.D. (2014). Degradation

kinetics and mechanism of β-lactam antibiotics by the activation of H2O2 and Na2S2O8

under UV-254 nm irradiation. Journal of Hazardous Materials 279, 375-383.

58. Heidelberger, C., Chaudhuri, N.K., Danneberg, P., Mooren, D., Griesbach, L., Duschinsky, R.,

et al. (1957). Fluorinated Pyrimidines, A New Class of Tumour-Inhibitory Compounds.

Nature 179(4561), 663-666.

59. Ho, P.L., Yung, R.W.H., Tsang, D.N.C., Que, T.L., Ho, M., Seto, W.H., et al. (2001). Increasing

resistance of Streptococcus pneumoniae to fluoroquinolones: results of a Hong Kong

multicentre study in 2000. Journal of Antimicrobial Chemotherapy 48(5), 659-665.

60. Hooper, D.C. (2002). Fluoroquinolone resistance among Gram-positive cocci. The Lancet

Infectious Diseases 2(9), 530-538.

61. Horemans, B., Vandermaesen, J., Vanhaecke, L., Smolders, E., and Springael, D. (2013).

Variovorax sp.-mediated biodegradation of the phenyl urea herbicide linuron at

micropollutant concentrations and effects of natural dissolved organic matter as

supplementary carbon source. Applied Microbiology and Biotechnology 97(22), 9837-9846.

63

62. Hu, J., Wang, W., Zhu, Z., Chang, H., Pan, F., and Lin, B. (2007). Quantitative

Structure−Activity Relationship Model for Prediction of Genotoxic Potential for Quinolone

Antibacterials. Environmental Science & Technology 41(13), 4806-4812.

63. Ihara, M., Satoh, K., Ishida, Y., Taniguchi, N., Tokunaga, Y., Takemura, M., et al. (1996). The

Synthesis of 1β-Fluorocarbapenem Derivative. Heterocycles 42(1), 437-443.

64. Isanbor, C., and O’Hagan, D. (2006). Fluorine in medicinal chemistry: A review of anti-

cancer agents. Journal of Fluorine Chemistry 127(3), 303-319.

65. Iwai, N., Sakai, R., Tsuchida, S., Kitazume, M., and Kitazume, T. (2009). Screening of

fluorinated materials degrading microbes. Journal of Fluorine Chemistry 130(4), 434-437.

66. Jaccaud, M., Faron, R., Devilliers, D., and Romano, R. (2000). "Fluorine," in Ullmann's

Encyclopedia of Industrial Chemistry. Wiley-VCH Verlag GmbH & Co. KGaA).

67. Jadhav, S.S., and David, M. (2016). Biodegradation of flubendiamide by a newly isolated

Chryseobacterium sp. strain SSJ1. 3 Biotech 6(1), 31.

68. Janssen, D.B., Oppentocht, J.E., and Poelarends, G.J. (2001). Microbial dehalogenation.

Current Opinion in Biotechnology 12(3), 254-258.

69. Jiang, M., Wang, L., and Ji, R. (2010). Biotic and abiotic degradation of four cephalosporin

antibiotics in a lake surface water and sediment. Chemosphere 80(11), 1399-1405.

70. Junker, T., Alexy, R., Knacker, T., and Kümmerer, K. (2006). Biodegradability of 14C-Labeled

Antibiotics in a Modified Laboratory Scale Sewage Treatment Plant at Environmentally

Relevant Concentrations. Environmental Science & Technology 40(1), 318-324.

71. Karigar, C.S., and Rao, S.S. (2011). Role of Microbial Enzymes in the Bioremediation of

Pollutants: A Review. Enzyme Research 2011, 11.

72. Karl, W., Schneider, J., and Wetzstein, H.-G. (2006). Outlines of an "exploding" network of

metabolites generated from the fluoroquinolone enrofloxacin by the brown rot fungus

Gloeophyllum striatum. Appl Microbiol Biotechnol 71.

73. Karlesa, A., De Vera, G.A.D., Dodd, M.C., Park, J., Espino, M.P.B., and Lee, Y. (2014).

Ferrate(VI) Oxidation of β-Lactam Antibiotics: Reaction Kinetics, Antibacterial Activity

Changes, and Transformation Products. Environmental Science & Technology 48(17),

10380-10389.

74. Kawasaki, H., Tsuda, K., Matsushita, I., and Tonomura, K. (1992). Lack of homology

between two haloacetate dehalogenase genes encoded on a plasmid from Moraxella sp.

strain B. Microbiology 138(7), 1317-1323.

75. Kawasaki, H., Yahara, H., and Tonomura, K. (1981). Isolation and Characterization of

Plasmid pUOl Mediating Dehalogenation of Haloacetate and Mercury Resistance in Moraxella

sp. B. Agricultural and Biological Chemistry 45(6), 1477-1481.

76. Kelly, M. (1965). Isolation of Bacteria able to metabolize Fluoroacetate or Fluoroacetamide.

Nature 208(5012), 809-810.

77. Key, B.D., Howell, R.D., and Criddle, C.S. (1997). Fluorinated Organics in the Biosphere.

Environmental Science & Technology 31(9), 2445-2454.

78. Kiel, M., and Engesser, K.-H. (2015). The biodegradation vs. biotransformation of

fluorosubstituted aromatics. Applied Microbiology and Biotechnology 99(18), 7433-7464.

79. Kim, B.R., Suidan, M.T., Wallington, T.J., and Du, X. (2000). Biodegradability of

Trifluoroacetic Acid. Environmental Engineering Science 17(6), 337-342.

80. Kirsch, P. (2013). "Introduction of Fluorine," in Modern Fluoroorganic Chemistry. Wiley-VCH

Verlag GmbH & Co. KGaA), 25-106.

81. Kumar, A., Bisht, B.S., Joshi, V.D., and Dhewa, T. (2011). Review on Bioremediation of

Polluted Environment: A Management Tool. International Journal of Environmental Sciences

1(6), 1079-1093.

82. Kurihara, T., and Esaki, N. (2008). Bacterial hydrolytic dehalogenases and related enzymes:

Occurrences, reaction mechanisms, and applications. The Chemical Record 8(2), 67-74.

83. Kurihara, T., Esaki, N., and Soda, K. (2000). Bacterial 2-haloacid dehalogenases: structures

and reaction mechanisms. Journal of Molecular Catalysis B: Enzymatic 10(1–3), 57-65.

84. Larsson, D.G.J., de Pedro, C., and Paxeus, N. (2007). Effluent from drug manufactures

contains extremely high levels of pharmaceuticals. Journal of Hazardous Materials 148(3),

751-755.

64

85. Li, W., Dai, Y., Xue, B., Li, Y., Peng, X., Zhang, J., et al. (2009). Biodegradation and

detoxification of endosulfan in aqueous medium and soil by Achromobacter xylosoxidans

strain CS5. Journal of Hazardous Materials 167(1–3), 209-216.

86. Li, X., Zheng, W., Machesky, M.L., Yates, S.R., and Katterhenry, M. (2011). Degradation

Kinetics and Mechanism of Antibiotic Ceftiofur in Recycled Water Derived from a Beef Farm.

Journal of Agricultural and Food Chemistry 59(18), 10176-10181.

87. Li, X.-W., Xie, Y.-F., Li, C.-L., Zhao, H.-N., Zhao, H., Wang, N., et al. (2014). Investigation of

residual fluoroquinolones in a soil–vegetable system in an intensive vegetable cultivation

area in Northern China. Science of The Total Environment 468–469, 258-264.

88. Liao, X., Li, B., Zou, R., Dai, Y., Xie, S., and Yuan, B. (2016). Biodegradation of antibiotic

ciprofloxacin: pathways, influential factors, and bacterial community structure.

Environmental Science and Pollution Research, 1-8.

89. Linderman, R.J., Jamois, E.A., and Tennyson, S.D. (1994). Synthesis of and Analysis of Thiol

Additions to .beta.-Alkyl-.alpha.,.beta.-unsaturated Trifluoromethyl Ketones. The Journal of

Organic Chemistry 59(5), 957-962.

90. Loi, E.I.H., Yeung, L.W.Y., Taniyasu, S., Lam, P.K.S., Kannan, K., and Yamashita, N. (2011).

Trophic Magnification of Poly- and Perfluorinated Compounds in a Subtropical Food Web.

Environmental Science & Technology 45(13), 5506-5513.

91. Loke, M.-L., Ingerslev, F., Halling-Sørensen, B., and Tjørnelund, J. (2000). Stability of Tylosin

A in manure containing test systems determined by high performance liquid

chromatography. Chemosphere 40(7), 759-765.

92. Maia, A.S., Ribeiro, A.R., Amorim, C.L., Barreiro, J.C., Cass, Q.B., Castro, P.M.L., et al. (2014).

Degradation of fluoroquinolone antibiotics and identification of metabolites/transformation

products by liquid chromatography–tandem mass spectrometry. Journal of Chromatography

A 1333, 87-98.

93. Marais, J.C.S. (1944). Monofluoroacetic acid, the toxic principle of "gifblaar" Dichapetalum

cymosum. Onderstepoort Jour. Vet. Sci. Animal Ind. 20(67).

94. Martens, R., Wetzstein, H.G., Zadrazil, F., Capelari, M., Hoffmann, P., and Schmeer, N.

(1996). Degradation of the fluoroquinolone enrofloxacin by wood-rotting fungi. Applied and

Environmental Microbiology 62(11), 4206-4209.

95. Martin, J.W., Franklin, J., Hanson, M.L., Solomon, K.R., Mabury, S.A., Ellis, D.A., et al. (2000).

Detection of Chlorodifluoroacetic Acid in Precipitation:  A Possible Product of Fluorocarbon

Degradation. Environmental Science & Technology 34(2), 274-281.

96. Martinez, J.L. (2009). Environmental pollution by antibiotics and by antibiotic resistance

determinants. Environmental Pollution 157(11), 2893-2902.

97. Meyer, J.J., Grobbelaar, N., and Steyn, P.L. (1990). Fluoroacetate-metabolizing pseudomonad

isolated from Dichapetalum cymosum. Applied and Environmental Microbiology 56(7),

2152-2155.

98. Miranda, C.D., and Castillo, G. (1998). Resistance to antibiotic and heavy metals of motile

aeromonads from Chilean freshwater. Science of The Total Environment 224(1–3), 167-176.

99. Morii, Y., Hasumi, F., and Kitazume, T. (2004). Accumulation of 2,2-difluoroethanol by

bacteria. Journal of Fluorine Chemistry 125(5), 731-733.

100. Murphy, C.D. (2016). Microbial degradation of fluorinated drugs: biochemical pathways,

impacts on the environment and potential applications. Applied Microbiology and

Biotechnology 100(6), 2617-2627.

101. Murphy, C.D., Clark, B.R., and Amadio, J. (2009). Metabolism of fluoroorganic compounds in

microorganisms: impacts for the environment and the production of fine chemicals. Applied

Microbiology and Biotechnology 84(4), 617-629.

102. Murray-Rust, P., Stallings, W.C., Monti, C.T., Preston, R.K., and Glusker, J.P. (1983).

Intermolecular interactions of the carbon-fluorine bond: the crystallographic environment of

fluorinated carboxylic acids and related structures. Journal of the American Chemical

Society 105(10), 3206-3214.

103. Neilson, A.H., and Allard, A.-S. (2002). "Degradation and Transformation of Organic Fluorine

Compounds," in Organofluorines, ed. A.H. Neilson. (Berlin, Heidelberg: Springer Berlin

Heidelberg), 137-202.

65

104. Ogilvie, S.C., Ataria, J.M., Waiwai, J., Doherty, J., Miller, A., Ross, J.G., et al. (2010).

Vertebrate pesticide risk assessment by indigenous communities in New Zealand.

Integrative Zoology 5(1), 37-43.

105. O'Hagan, D. (2008). Understanding organofluorine chemistry. An introduction to the C-F

bond. Chemical Society Reviews 37(2), 308-319.

106. O'Hagan, D., Perry, R., Lock, J.M., Meyer, J.J.M., Dasaradhi, L., Hamilton, J.T.G., et al. (1993).

High levels of monofluoroacetate in Dichapetalum braunii. Phytochemistry 33(5), 1043-

1045.

107. O'Halloran, K., Jones, D., Booth, L., and Fisher, P. (2005). Ecotoxicity of sodium

fluoroacetate (compound 1080) to soil organisms. Environmental Toxicology and Chemistry

24(5), 1211-1218.

108. Ojima, I. (2013). Exploration of Fluorine Chemistry at the Multidisciplinary Interface of

Chemistry and Biology. The Journal of organic chemistry 78(13), 6358-6383.

109. Okazoe, T. (2009). Overview on the history of organofluorine chemistry from the viewpoint

of material industry. Proceedings of the Japan Academy. Series B, Physical and Biological

Sciences 85(8), 276-289.

110. Oliveira, R., Dodd, J., and Castro, P. (2001). The mycorrhizal status of Phragmites australis

in several polluted soils and sediments of an industrialised region of Northern Portugal.

Mycorrhiza 10(5), 241-247.

111. Pan, X., Lin, D., Zheng, Y., Zhang, Q., Yin, Y., Cai, L., et al. (2016). Biodegradation of DDT by

Stenotrophomonas sp. DDT-1: Characterization and genome functional analysis. Scientific

Reports 6, 21332.

112. Parshikov, I.A., Freeman, J.P., Lay, J.O., Beger, R.D., Williams, A.J., and Sutherland, J.B.

(1999). Regioselective transformation of ciprofloxacin to N-acetylciprofloxacin by the

fungus Mucor ramannianus. FEMS Microbiology Letters 177(1), 131-135.

113. Parshikov, I.A., Freeman, J.P., Lay, J.O., Beger, R.D., Williams, A.J., and Sutherland, J.B.

(2000). Microbiological transformation of enrofloxacin by the fungus Mucor ramannianus.

Appl Environ Microbiol 66.

114. Percy, J.M. (1997). "Building Block Approaches to Aliphatic Organofluorine Compounds," in

Organofluorine Chemistry: Techniques and Synthons, ed. R.D. Chambers. (Berlin,

Heidelberg: Springer Berlin Heidelberg), 131-195.

115. Picó, Y., and Andreu, V. (2006). Fluoroquinolones in soil—risks and challenges. Analytical

and Bioanalytical Chemistry 387(4), 1287-1299.

116. Piekarz, A.M., Primbs, T., Field, J.A., Barofsky, D.F., and Simonich, S. (2007). Semivolatile

Fluorinated Organic Compounds in Asian and Western U.S. Air Masses. Environmental

Science & Technology 41(24), 8248-8255.

117. Prasad, B., and Suresh, S. (2012). Biodegradation of Phthalate Esters by Variovorax sp.

APCBEE Procedia 1, 16-21.

118. Prasad, B., and Suresh, S. (2015). Biodegradation of dimethyl phthalate ester using free

cells, entrapped cells of Variovorax sp. BS1 and cell free enzyme extracts: A comparative

study. International Biodeterioration & Biodegradation 97, 179-187.

119. Rafii, F., Williams, A.J., Park, M., Sims, L.M., Heinze, T.M., Cerniglia, C.E., et al. (2009).

Isolation of bacterial strains from bovine fecal microflora capable of degradation of

ceftiofur. Veterinary Microbiology 139(1–2), 89-96.

120. Reis, P.J.M., Reis, A.C., Ricken, B., Kolvenbach, B.A., Manaia, C.M., Corvini, P.F.X., et al.

(2014). Biodegradation of sulfamethoxazole and other sulfonamides by Achromobacter

denitrificans PR1. Journal of Hazardous Materials 280, 741-749.

121. Rex, E.H., and Susan, F.K. (2002). Cephalosporins in Veterinary Medicine - Ceftiofur Use in

Food Animals. Current Topics in Medicinal Chemistry 2(7), 717-731.

122. Römpp, A., Klemm, O., Fricke, W., and Frank, H. (2001). Haloacetates in Fog and Rain.

Environmental Science & Technology 35(7), 1294-1298.

123. Sambrook, J., Fritcsh, E.F., and Maniatis, T. (1989). Molecular cloning: a laboratory manual.

New York: Cold Spring Harbor Laboratory Press.

124. Sanada, M., Miyano, T., Iwadare, S., Williamson, J.M., Arison, B.H., Smith, J.L., et al. (1986).

Biosynthesis of fluorothreonine and fluoroacetic acid by the thienamycin producer,

Streptomyces cattleya. The Journal of Antibiotics 39(2), 259-265.

66

125. Sarmah, A.K., Meyer, M.T., and Boxall, A.B.A. (2006). A global perspective on the use, sales,

exposure pathways, occurrence, fate and effects of veterinary antibiotics (VAs) in the

environment. Chemosphere 65(5), 725-759.

126. Satola, B., Wübbeler, J.H., and Steinbüchel, A. (2013). Metabolic characteristics of the

species Variovorax paradoxus. Applied Microbiology and Biotechnology 97(2), 541-560.

127. Scott, B.F., Macdonald, R.W., Kannan, K., Fisk, A., Witter, A., Yamashita, N., et al. (2005).

Trifluoroacetate Profiles in the Arctic, Atlantic, and Pacific Oceans. Environmental Science &

Technology 39(17), 6555-6560.

128. Shannon, M.J.R., and Unterman, R. (1993). Evaluating Bioremediation: Distinguishing Fact

from Fiction. Annual Review of Microbiology 47(1), 715-736.

129. Siegemund, G., Schwertfeger, W., Feiring, A., Smart, B., Behr, F., Vogel, H., et al. (2000).

"Fluorine Compounds, Organic," in Ullmann's Encyclopedia of Industrial Chemistry. Wiley-

VCH Verlag GmbH & Co. KGaA).

130. Sim, W.-J., Lee, J.-W., Lee, E.-S., Shin, S.-K., Hwang, S.-R., and Oh, J.-E. (2011). Occurrence

and distribution of pharmaceuticals in wastewater from households, livestock farms,

hospitals and pharmaceutical manufactures. Chemosphere 82(2), 179-186.

131. Smit, M.F., van Heerden, P.D.R., Pienaar, J.J., Weissflog, L., Strasser, R.J., and Krüger, G.H.J.

(2009). Effect of trifluoroacetate, a persistent degradation product of fluorinated

hydrocarbons, on Phaseolus vulgaris and Zea mays. Plant Physiology and Biochemistry

47(7), 623-634.

132. Sota, M., Endo, M., Nitta, K., Kawasaki, H., and Tsuda, M. (2002). Characterization of a Class

II Defective Transposon Carrying Two Haloacetate Dehalogenase Genes from Delftia

acidovorans Plasmid pUO1. Applied and Environmental Microbiology 68(5), 2307-2315.

133. Spargo, P.L. (2005). Handbook of Fluorous Chemistry. American Chemical Society.

134. Sturini, M., Speltini, A., Maraschi, F., Profumo, A., Pretali, L., Irastorza, E.A., et al. (2012).

Photolytic and photocatalytic degradation of fluoroquinolones in untreated river water

under natural sunlight. Applied Catalysis B: Environmental 119–120, 32-39.

135. Su, L.-H., Chu, C., Cloeckaert, A., and Chiu, C.-H. (2008). An epidemic of plasmids?

Dissemination of extended-spectrum cephalosporinases among Salmonella and other

Enterobacteriaceae. FEMS Immunology & Medical Microbiology 52(2), 155-168.

136. Sylvestre, M. (1995). Biosorption and BioremediationBiphenyl/Chlorobiphenyls catabolic

pathway of Comamonas testosteroni B-356: Prospect for use in bioremediation.

International Biodeterioration & Biodegradation 35(1), 189-211.

137. Tamura, K., Mizukami, H., Maeda, K., Watanabe, H., and Uneyama, K. (1993). One-pot

synthesis of trifluoroacetimidoyl halides. The Journal of Organic Chemistry 58(1), 32-35.

138. Tasho, R.P., and Cho, J.Y. (2016). Veterinary antibiotics in animal waste, its distribution in

soil and uptake by plants: A review. Science of The Total Environment 563–564, 366-376.

139. Thayer, A.M. 2006. Fabulous Fluorine. Chemical & Engineering News [Online], 84. Available:

http://pubs.acs.org/email/cen/html/060606071341.html. [Accessed 18/07/2016]

140. Thompson, I.P., Van Der Gast, C.J., Ciric, L., and Singer, A.C. (2005). Bioaugmentation for

bioremediation: the challenge of strain selection. Environmental Microbiology 7(7), 909-

915.

141. Uneyama, K. (2007). "Fundamentals in Organic Fluorine Chemistry," in Organofluorine

Chemistry. Blackwell Publishing), 1-100.

142. US-EPA (2007). "Treatment Technologies for Mercury in Soil, Waste and Water". Office of

Superfund Remediation and Technology Innovation.

143. Visscher, P.T., Culbertson, C.W., and Oremland, R.S. (1994). Degradation of trifluoroacetate

in oxic and anoxic sediments. Nature 369(6483), 729-731.

144. Walsh, C. (2000). Molecular mechanisms that confer antibacterial drug resistance. Nature

406(6797), 775-781.

145. Wang, F., Dicinoski, G.W., Zhu, Y., and Haddad, P.R. (2004). Simultaneous determination of

monofluoroacetate, difluoroacetate and trifluoroacetate in environmental samples by ion

chromatography. Journal of Chromatography A 1032(1–2), 31-35.

146. Wang, J., and Chen, C. (2009). Biosorbents for heavy metals removal and their future.

Biotechnology Advances 27(2), 195-226.

147. Weisburg, W.G., Barns, S.M., Pelletier, D.A., and Lane, D.J. (1991). 16S ribosomal DNA

amplification for phylogenetic study. Journal of Bacteriology 173(2), 697-703.

67

148. Weist, K., Muller, A., Monnet, D., and Heuer, O. (2014). "Surveillance of antimicrobial

consumption in Europe". European Centre for Disease Prevention and Control).

149. Wetzstein, H.-G., Schmeer, N., and Karl, W. (1997). Degradation of the fluoroquinolone

enrofloxacin by the brown rot fungus Gloeophyllum striatum: identification of metabolites.

Appl Environ Microbiol 63.

150. Wetzstein, H.-G., Schneider, J., and Karl, W. (2006). Patterns of metabolites produced from

the fluoroquinolone enrofloxacin by basidiomycetes indigenous to agricultural sites. Appl

Microbiol Biotechnol 71.

151. Wetzstein, H.-G., Schneider, J., and Karl, W. (2009). "Comparative biotransformation of

fluoroquinolone antibiotics in matrices of agricultural relevance," in Veterinary

pharmaceuticals in the environment. ACS Symposium Series, eds. K.L. Henderson & J.R.

Coats. (Washington, DC: American Chemical Society).

152. Wong, D.H., Kirkpatrick, W.E., King, D.R., and Kinnear, J.E. (1992). Defluorination of sodium

monofluoroacetate (1080) by microorganisms isolated from western Australian soils. Soil

Biology and Biochemistry 24(9), 833-838.

153. Xu, C., Ding, J., Qiu, J., and Ma, Y. (2013). Biodegradation of acetochlor by a newly isolated

Achromobacter sp. strain D-12. Journal of Environmental Science and Health, Part B 48(11),

960-966.

154. Yamazaki, T., Taguchi, T., and Ojima, I. (2009). "Unique Properties of Fluorine and their

Relevance to Medicinal Chemistry and Chemical Biology," in Fluorine in Medicinal Chemistry

and Chemical Biology. John Wiley & Sons, Ltd), 1-46.

155. Yang, B., Zuo, J., Li, P., Wang, K., Yu, X., and Zhang, M. (2016). Effective ultrasound

electrochemical degradation of biological toxicity and refractory cephalosporin

pharmaceutical wastewater. Chemical Engineering Journal 287, 30-37.

156. Zamanpour, G., and Mehrabani-Zeinabad, A. (2014). An experimental study on

bioremediation and photolysis of enrofloxacin. Water Science & Technology 70(5), 932-938.

157. Zhang, T., and Li, B. (2011). Occurrence, Transformation, and Fate of Antibiotics in

Municipal Wastewater Treatment Plants. Critical Reviews in Environmental Science and

Technology 41(11), 951-998.

158. Zhang, X.-J., Lai, T.-B., and Kong, R.Y.-C. (2012). "Biology of Fluoro-Organic Compounds," in

Fluorous Chemistry, ed. T.I. Horváth. (Berlin, Heidelberg: Springer Berlin Heidelberg), 365-

404.

159. Zhao, L., Dong, Y.H., and Wang, H. (2010). Residues of veterinary antibiotics in manures

from feedlot livestock in eight provinces of China. Science of The Total Environment 408(5),

1069-1075.